Environmental Toxicology and Chemistry, Vol. 24, No. 8, pp. 2037–2044, 2005
? 2005 SETAC
Printed in the USA
0730-7268/05 $12.00 ? .00
EFFECT OF COEXPOSURE TO DDT AND MANGANESE ON FRESHWATER
INVERTEBRATES: PORE WATER FROM CONTAMINATED RIVERS AND
JESU´S MEJI´A-SAAVEDRA,*† SERGIO SA´NCHEZ-ARMASS,† GUSTAVO E. SANTOS-MEDRANO,‡
ROBERTO GONZA´LEZ-AMARO,† ISRAEL RAZO-SOTO,† ROBERTO RICO-MARTI´NEZ,‡ and FERNANDO DI´AZ-BARRIGA†
†Facultad de Medicina, Universidad Auto ´noma de San Luis Potosı ´, San Luis Potosı ´, SLP 78210, Mexico
‡Departamento de Quı ´mica, Centro de Ciencias Ba ´sicas, Universidad Auto ´noma de Aguascalientes, Aguascalientes, Ags 20110, Mexico
(Received 17 August 2004; Accepted 15 February 2005)
Abstract—An environmental survey of several rivers of the southern Huasteca area of Mexico revealed high concentrations of
manganese (Mn) and the presence of DDT in the sediments and pore water. Therefore, acute (48-h) toxicities of Mn and DDT
were assessed both independently and as a combination on 24-h-old neonates of Daphnia magna Strauss and Lecane quadridentata
Ehrenberg. Daphnia magna showed high sensitivity to both toxicants, whereas L. quadridentata was highly resistant to DDT and
less susceptible to Mn. For D. magna, the Mn and DDT coexposure was significantly more toxic than any of the singly tested
compounds. When D. magna was exposed to sediment pore water, no association was found between the Mn content in the samples
and the observed toxicity. Preliminary particle analysis of pore water showed different compounds of Mn, which apparently were
not in bioavailable form.
Dichlorodiphenylthrichloroethane Invertebrate acute toxicity Toxicant coexposurePore
The toxicity assessment of chemical combinations is a
growing necessity, because in the environment of hazardous
waste sites, single chemicals rarely are found. However, most
toxicological studies are performed using experimentaldesigns
with only one chemical. In fact, 95% of the money spent in
toxicological studies is still assigned to individual chemicals
. Coexposure to toxic agents may have different toxicolog-
ical characteristics compared with the biological effects of
single compounds or elements [2–4].
Pesticides and metals have been found in streams, where
they bioaccumulate in aquatic organisms, causing adverse ef-
fects [5–7]. Still, information about the biological effects of
combinations of pesticides and metals is scarce. This is the
case for manganese (Mn) and DDT, two environmental con-
taminants that have been studied previously in individual form
[8,9] but never, to our knowledge, in combination.
In Mexico, two scenarios exist in which coexposure to Mn
and DDT may occur. The first occurs in the state of San Luis
Potosı ´, where the Mn source is a mine that discharges mining
wastes into the regional aquatic ecosystem. Data collected dur-
ing the early part of the present study found areas with high
Mn concentrations in sediments at this location. In addition,
DDT has been used in this area in the control program for
malaria from the year 1957 to the year 2000. The second
scenario for coexposure of DDT and Mn occurs in southeastern
Mexico, specifically in the state of Chiapas. In this area, the
source of Mn was the widespread use of the fungicide maneb
(ethylen-bis-dithiocarbamate of Mn), which has been used in
banana plantations, whereas the source of DDT, as in the case
of San Luis Potosı ´, was the program against malaria. Taking
* To whom correspondence may be addressed (email@example.com).
into account that both DDT and Mn can be found in streams
in San Luis Potosı ´ and Chiapas, the ecotoxicity of this com-
bination needs to be assessed.
Freshwater invertebrates in general, and daphnids in par-
ticular, have been widely used in toxicological studies of water
reservoirs contaminated with anthropogenic toxics [10–12].
Use of invertebrates in the assessment of aquatic ecosystems
is justified given the importance of these organisms as primary
consumers; furthermore, such organisms have been used in the
evaluation of different responses, such as lethality , en-
docrine disruption [14–16], enzymatic inhibition , and
feeding-rate inhibition .
In the present study, we determined environmental con-
centrations of DDT and Mn at seven stations located in the
southern Huasteca area of Mexico. On confirming the presence
of elevated concentrations of both DDT and Mn (mainly in
the sediments), we decided to study the effects of DDT and
Mn coexposure using, as experimental models, two inverte-
brates representative of two different phyla (Rotifera and Ar-
thropoda) of the aquatic biota: The crustacean Daphnia magna
(a truly planktonic species), and the rotifer Lecane quadri-
dentata (a littoral species often found buried in the sediments).
MATERIALS AND METHODS
We used the D. magna Strauss acute toxicity protocol de-
ly, this technique consists of 48-h exposure of D. magna ne-
onates less than 24 h old. Control (U.S. Environmental Pro-
tection Agency [U.S. EPA] medium) and treatment groups
(EPA medium with 0.0085, 0.017, 0.034, 0.068, 0.135, 0.271,
or 0.406 ?M DDT; EPA medium with 4.6, 9.1, 18.2, 72.8,
109.2, or 145.6 ?M Mn) were studied. The DDT was dissolved
Environ. Toxicol. Chem. 24, 2005J. Mejı ´a-Saavedra et al.
Fig. 1. Map of the southern Huasteca area around the city of Tama-
zunchale, San Luis Potosi, Mexico, showing the sampling stations.
The Salado and Huitznopala rivers are influents of the Claro River.
Tamala and Tenexco are sites located on the banks of the Claro River.
Tancuilin, Axtla, and Huichihuayan are sampling stations on the rivers
of the same name.
in acetone. In the control group, 10 neonates were placed in
each of three glass beakers (this set of beakers represents one
replicate) with 100 ml of EPA medium (EPA medium consisted
of 96 mg of NaHCO3, 60 mg of CaSO4·2H2O, 123 mg of
MgSO4·7H2O, and 4 mg of KCl per liter of deionized water)
. The solution pH was adjusted to 7.5. No food was given
to the neonates during the test period (48 h), and the organisms
were kept in an incubator with a 18:6-h light:dark schedule.
Light intensity fluctuated between 600 and 1,100 lux (in any
part of the chamber) as determined by an illuminometer (Kyor-
itan Electrical Instruments, Tokyo, Japan). Temperature was
kept at 20 ? 1?C as determined by a microcomputer ther-
mometer (Hanna Instruments, Villafranca, Italy). Lack of
movement was the end point used to determine the animals
as dead in five replicate batches of D. magna. All chemicals
were of the highest available purity (for Mn, Mn atomic ab-
sorption standard solution; for DDT, 1,1-bis(p-chlorophenyl)-
2,2,2-trichloroethane; Sigma, St. Louis, MO, USA). Nominal
concentrations were used to calculate the median lethal con-
centration (LC50). The combinations of DDT and Mn were
assessed using a concentration corresponding to the no-ob-
served-effect concentration (NOEC) of one agent plusthe same
range of concentration previously tested for the other toxicant.
For L. quadridentata Ehrenberg, parthenogenetic eggs were
collected 24 h before the start of experiments. Five-hundred
eggs were placed in a 1-ml well with EPA medium at pH 7.5.
The eggs were incubated in a controlled temperature chamber
(Revco, Asheville, NC, USA) with a 16:8-h light:darkschedule
at 25?C. Hatching efficiency (24 h) was always greater than
37% under these conditions. Neonates less than 24 h old were
then collected, and we used the protocol developed by Pe ´rez-
Legaspi and Rico-Martı ´nez . The test was begun by pi-
petting 1 ml of test solution (DDT: 0.018, 0.035, 0.071, 0.141,
0.28, 0.56, or 1.41 ?M; Mn: 4.6, 18.2, 36.4, 91.0, or 182.1
?M) or control EPA medium into a 24-well plate. Because
acetone was required as a solvent for DDT, the control was
exposed to acetone 0.5 M. Ten animals were transferred into
the test well with a plastic transfer micropipette. Transfer was
done under stereoscopic microscope at ?10 magnification.
Five replicates were used for each treatment.
Statistical analyses were performed with the R software
(http://www.r-project.org/) at 95% confidence level. To obtain
variance homogeneity (Fligner–Killen test) and normality
(Shapiro–Wilk test) of the residuals, the LC50 values were log
transformed before the statistical analysis. One-way analysis
of variance was applied to the data. Multiple comparisons were
performed with Tukey’s honest-significance-difference test.
The LC50 values were determined with the program DL-50
(S.B.I.-I.R.C.T., Montpellier, France), which tests the validity
of the linear fitting by using chi-square statistics (p ? 0.05).
Geographical area of sampling
Four rivers were included in the region selected for the
present study. In this region, DDT was used in the malaria-
control program. However, whereas the Claro River receives
discharges from a Mn mining area, the Tancuilin, the Hui-
chihuayan, and the Axtla rivers are not impacted by this mining
area (Fig. 1). Seven sampling stations were identified, four of
them located in the Claro River: One upriver from the site
where the river receives the waters coming from the mining
area (Salado), another located in this site (Huitznopala), and
two more downriver, one situated 10 km away (Tamala) and
the other 20 km away (Tenexco). The rest of the sampling
stations were located in the other three rivers (one in each
river). In each station, we took three samples spaced 0.5 km
Pore water and sediment samples (surface sediment; depth,
0–5 cm) were collected in November 2001. Sediments were
collected with a flat-hand shovel, and the pore water was ex-
tracted on site by suction using a ceramic air disperser adapted
to a syringe. Sediment was collected in polypropylene bags.
For metal analysis, water was collected in polypropylene bot-
tles previously rinsed with 10% HNO3. For DDT analysis,
sediment and pore water were collected and transported to the
laboratory in glass containers and kept under refrigeration
(4?C) until analysis.
Analysis of toxicant concentrations
Manganese analysis was performed with a 2380 atomic
absorption spectrophotometer (Perkin-Elmer, Boston, MA,
USA). Analysis of standard reference material was conducted
as an internal quality control in each run. For sediment, Stan-
dard Reference Material (SRM) 2704 (Buffalo River sediment;
National Institute of Standards and Technology, Gaithersburg,
MD, USA) was used, with recoveries of 93%. For water, SRM
1643d (trace elements in water) was used, with recoveries of
91%. Each sample was analyzed in duplicate. Deionized water
was used for all analytical work, and glassware and other
materials were soaked in 10% nitric acid, rinsed with deionized
Coexposure of DDT and manganese
Environ. Toxicol. Chem. 24, 20052039
Table 1. Results of the acute tests performed with Daphnia magnaa
aAll results are expressed as ?M. LC50 ? median lethal concentra-
tion; LOEC ? lowest-observed-effect concentration; NOEC ? no-
Fig. 2. Comparison of Daphnia magna lethality exposed to DDT, Mn,
and a mixture of both toxicants. (a) Daphnia magna exposed to DDT
and DDT/Mn. (b) Daphnia magna exposed to Mn and Mn/DDT. The
dotted lines correspond to the regression line of the toxicityinteraction
for both toxicants. The continuous lines correspond to the regression
line of the toxicity of each individual toxicant. Data represent the
mean ? 1 standard deviation.
Table 2. Results of acute tests performed with Lecane
aAll results are expressed as ?M. No effects were observed within
the evaluated DDT concentration range (0.017–1.4 ?M). LC50 ?
median lethal concentration; LOEC ? lowest-observed-effect con-
centration; ND ? not determined; NOEC ? no-observed-effect con-
water, and dried before use. The DDT in water was analyzed
according to the U.S. EPA method (http://www.speclab.com/
compound/m508.htm). The DDT in sediments was analyzed
as described by Yan ˜ez et al. . Briefly, sediment samples
(1 g) were microwave extracted in 15 ml of methylenechlorine.
After the extraction, samples were evaporated to 0.2 ml by
nitrogen current, and they were resuspended to 2.0 ml with
hexane. This procedure was repeated five times. The last re-
suspension was to 6.0 ml, and this was transferred to a florisil
column where the extraction was done with 6% ethyl ether in
hexane. The florisil eluate was concentrated to 1 ml by nitrogen
current. Quantitative analyses were performed by gas chro-
matography using a Varian model 3400 (Palo Alto, CA, USA)
equipped with electron-capture detector.
The size of particles in pore-water samples was determined
with laser-ray diffraction (Sald 1100; Shimadzu, Kyoto, Japan)
and analyzed with scanning electron microscopy (ZL30; Phil-
lips, Eindhoven, The Netherlands) equipped with an energy-
dispersive x-ray spectrophotometer (DX40; EDAX, Mahwah,
NJ, USA) to obtain chemical composition of particles. Eight
to 12 particles were analyzed from three samples: Tamala 3,
Axtla 3, and Tancuilin 3.
Daphnia magna. Considering the LC50 values (Table 1),
this organism clearly was approximately 700-fold more sen-
sitive to DDT than to Mn (p ? 0.05). Furthermore, the com-
bination DDT/Mn was 3.7-fold more toxic than DDT (p ?
0.05) (Fig. 2a), whereas the combination Mn/DDT was eight-
fold more toxic than Mn alone (p ? 0.05); DDT/Mncorrespond
to the range of concentrations used for DDT and the no-ob-
served-effect concentration (NOEC) for Mn. In addition, Mn/
DDT correspond to the range of concentrations used for Mn
and the NOEC for DDT (Fig. 2b). Therefore, Mn increased
the toxicity of DDT by 3.7-fold, and this insecticide increased
the toxicity of the metal by eightfold. It is worthwhile to men-
tion the interaction between Mn and DDT shown by the lack
of parallelism between the regression lines of the probit plot.
It was observed that in the Mn/DDT mixture, the additional
toxicity of the presence of DDT at its NOEC level became
less as the Mn concentration increased (Fig. 2a). The same
behavior was observed with the combination DDT/Mn (Fig.
Lecane quadridentata. The results of LC50 values of L.
quadridentata are shown in Table 2. The response of L. quad-
ridentata contrasted sharply with the response of D. magna.
This rotifer was extremely resistant to DDT (highest concen-
tration tested, 1.41 ?M) and to the DDT/Mn coexposure. Fur-
thermore, it was twofold less susceptible to Mn compared with
D. magna (p ? 0.05) (Tables 1 and 2). When L. quadridentata
was exposed to the Mn/DDT combination, an apparent dim-
inution in toxicity was observed with respect to Mn (Table 2).
However, in the coexposure assays, we did not reach 100%
lethality, and the 95% confidence limits of the LC50 were very
Environ. Toxicol. Chem. 24, 2005 J. Mejı ´a-Saavedra et al.
Table 3. Environmental concentration of Mn and DDT in the sampling stations in San Luis
18.9 ? 2.8
966.5 ? 246.0
201.8 ? 80.2
270.8 ? 38.0
10.4 ? 0.95
8.8 ? 0.62
10.6 ? 0.91
1.1 ? 1.4
52.8 ? 15.4
7.3 ? 2.3
4.1 ? 2.6
9.2 ? 6.4
5.1 ? 5.2
1.9 ? 2.1
22.6 ? 8.5
36.7 ? 22.6
62.1 ? 5.6
141.0 ? 11.3
56.4 ? 22.6
aThe concentration of DDT in the pore water was not detectable (ND). The DDT detection limit was
0.0016 ?M. Values correspond to the mean ? standard deviation (n ? 3, except for Salado and
Huitznopala sediment, where n ? 1).
bNational Ocean and Atmospheric Administration Screening quick references tables .
cNormal Official Mexicana .
dCanadian environmental quality guidelines .
Table 4. Toxicity tests of pore water with Daphnia magnaa
4.05 ? 10?5
aThe numbers 1, 2, and 3 correspond to collection sites in each river in San Luis Potosi, Mexico.
Expected lethality was based on individual Mn concentration. The p-values correspond to one-tailed
t tests between the observed and expected lethality (n ? 3, except for Huichihuayan, where n ? 2).
broad, overlapping with the confidence limits of the Mn LC50,
indicating no significant differences (Table 2).
Mn and DDT concentrations in the screened rivers
Levels of Mn and DDT in sediments and pore water are
shown in Table 3. As expected, Mn levels in the Claro stations
were higher than in Huitznopala, and they decrease as the
distance from the mine discharge increases. Interestingly, the
levels of Mn in pore water in the Tancuilin station were as
high as those found in the Tamala station (in the Claro River),
whereas the levels in the Huichihuayan station were similar
to those found in a nonpolluted station (Salado). With regard
to DDT, it is important to note that the insecticide was not
detected (detection limit, 0.0016 ?M) in any of the pore-water
samples (Table 3). In contrast, in the sediment samples, we
found that whereas at the Salado station, DDT levels (15.2
?mol/kg) were less than the Canadian Sediment Quality
Guideline for the Protection of Aquatic Life of 17.5 ?mol/kg
(http://www.ec.gc.ca), at the Tamala (22.6 ?mol/kg), Tancuilin
(62.1 ?mol/kg), and Axtla (141.0 ?mol/kg) stations, the levels
were higher than this guideline (Table 3).
Pore-water toxicity test
Because the D. magna bioassay presented a better response
in the laboratory test than the L. quadridentata bioassay, we
used the cladoceran for the pore-water toxicity assessment.
The results did not show a relation between lethality and the
Mn concentration in the samples (p ? 0.8982). Table 4 shows
the results obtained when D. magna was exposed to pore water.
The expected toxicity values were based on the Mn concen-
trations in the pore water. The lethality observed in the Huitz-
nopala and Tamala stations was lower than expected (Table
4). Besides, overall analysis of the observed versus expected
toxicity showed significant differences (p ? 0.05).
Coexposure of DDT and manganese
Environ. Toxicol. Chem. 24, 20052041
Fig. 3. Results of the scanning electron microscopy, electronic micrography, energy-dispersive spectrum, and element weight percentage (Wt%)
analyses of two typical particles from Tamala 3 pore water. Tamala is a site located on the banks of the Claro River in San Luis Potosi, Mexico.
(a) A particle rich in Mn oxide. (b) A Fe-Mn oxyhydroxide particle.
The results in the preceding section suggested that Mn was
not bioavailable. For particle analysis, 8 to 12 particles from
each of three samples with the same Mn content were selected
(Tamala 3, Tancuilin 3, and Axtla 3). The particle size and the
elements contained in the particles were determined. No dif-
ferences were observed among the three samples with respect
to particle size, with all samples presenting particles less than
40 ?m in diameter. With respect to Mn compounds, the Tamala
3 sample had Mn oxide particles (Fig. 3a) and iron-Mn oxy-
hydroxides (Fig. 3b). The main elements in the Tamala 3 par-
ticles (n ? 12) were Mn (42.1% ? 25.8%), O (21.9% ? 5.6%),
C (16.2% ? 17.3%), and Fe (12.0% ? 15.5). The Axtla 3
sample (Fig. 4) had particles (n ? 8) rich in C (42.5% ?
8.2%), O (26.5% ? 3.9%), and Mn (15.0% ? 8.4%). The
Tancuilin 3 sample (Fig. 5) had particles (n ? 8) with O (31.9%
? 3.5%), C (26.8% ? 4.1%), and Mn (26.3% ? 4.1%). Re-
garding the physicochemical characterization of the water (Ta-
ble 5), no clear differences were observed.
To our knowledge, we present the highest environmental
concentration of Mn reported for sediments and provide the
first report where DDT and Mn coexposure was evaluated
using two experimental organisms representing different phyla
We observed a differential response in both organisms to
the lethal effects of either DDT or Mn. Lecane quadridentata
was highly resistant to DDT exposure and less susceptible to
Mn than D. magna (Tables 1 and 2). Considering DDT, L.
quadridentata did not show any lethal effect at concentrations
as high as 1.41 ?M (0.5 mg/L), the highest DDT concentration
used in the present study to avoid toxic effect of the solvent
(acetone), because the control group presented elevated mor-
tality from increasing acetone concentration.
Serrano et al.  reported that another rotifer, Brachionus
plicatilis, is approximately 1,000-fold more resistant to pes-
ticides compared with other aquatic organisms, such as am-
phipods  or fishes . Rotifers also are resistant to other
pesticides, such as lindane and trichlorfon , malathion and
diazinon , and fenitrothion . The resistance to DDT is
not a peculiarity of rotifers, because Hydra attenuata also is
resistant to DDT .
In regard to the results obtained with D. magna, the LC50
value for DDT reported in the present study (0.024 ?M) is
similar to the 0.017 ?M reported by Rawash et al.  and
is higher than the 0.0085 ?M reported by Ziegenfuss et al.
, whereas the Mn LC50 reported here (17.6 ?M) was lower
than the 150.7 ?M reported by Khangarot and Ray . It is
Environ. Toxicol. Chem. 24, 2005J. Mejı ´a-Saavedra et al.
Fig. 4. Results of the scanning electron microscopy, electronic micrography, energy-dispersive spectrum, and element weight percentage (Wt%)
analyses of a typical particle from Axtla 3 pore water. Axtla is a sampling station on the river of the same name in San Luis Potosi, Mexico.
Fig. 5. Results of the scanning electron microscopy, electronic micrography, energy-dispersive spectrum, and element weight percentage (Wt%)
analyses of a typical particle from Tancuilin 3 pore water. Tancuilin is a sampling station on the river of the same name in San Luis Potosi,
worth mentioning that differences within one order of mag-
nitude among D. magna toxicity tests (using a similarprotocol)
are considered to be acceptable.
The coexposure results obtained with DDT and Mn in terms
of the NOEC were relevant. In D. magna, both combinations
(DDT/Mn and Mn/DDT) (Table 1) were more toxic than DDT
and Mn alone. The coexposure lethality mechanism is un-
known. However, these results are important, because in the
environment, the toxicants are mixed and, in many cases, occur
at low concentrations that may be regarded as nonadverse. Our
results showed that one toxicant, at the NOEC concentration,
can increase the toxicity of the other. The mechanism that
underlies the toxicological interaction observed in the coex-
posure (Mn/DDT and DDT/Mn) requires further investigation.
Because of the results of the toxicity tests, D. magna was
used in assessment of the toxicity of pore water from the rivers
of interest. For this purpose, interstitial water (pore water) was
collected, because we expected that the levels of DDT and Mn
in this media would be related to the levels found in sediments.
However, this was not the case for DDT. The concentrations
of the insecticide were not detectable in any of the pore-water
samples, although DDT levels in sediments were higher than
the guidelines. Therefore, our results using D. magna referred
only to the Mn content.
As shown in Table 4, observed mortality did not correlate
with the Mn concentration measured in pore-water samples.
The pore-water toxicity generally was lower than the expected
toxicity, and this was particularly evident for the samples col-
lected in the Claro River. A possible explanation for this result
is that the Mn at these stations was in a nonbioavailable form.
In fact, we found Mn oxides (Fig. 3a) and iron-Mn oxyhy-
droxides (Fig. 3b), which are very stable compounds consid-
ering the physicochemical conditions in the rivers (Table 5).
In the other stations (i.e., the Axtla and Tancuilin rivers), Mn
may be adsorbed to organic materials (Figs. 4 and 5). Sorption
of metal ions usually reduces their concentration in water and,
in many cases, their bioavailability as well . In turn, Wel-
tens et al. , in a bioaccumulation experiment with D. mag-
Coexposure of DDT and manganese
Environ. Toxicol. Chem. 24, 2005 2043
Table 5. Physicochemical parameters of water samplesa
Tamala 3 Tancuilin 3 Axtla 3
Alkalinity (mg CaCO3/L)
Hardness (mg CaCO3/L)
Total dissolved solids (mg/L)
Oxygen dissolved (mg/L)
aTamala is a site located on the banks of the Claro River in San Luis
Potosi, Mexico. Axtla is a sampling station on the river of the same
name in San Luis Potosi, Mexico. Tancuilin is a sampling station
on the river of the same name in San Luis Potosi, Mexico. The
number 3 corresponds to the collection site in each river.
na, showed that metals like cadmium and zinc adsorbed to
organic or mineral materials are bioavailable to the cladoceran.
Those authors mention that desorption in the gut probably is
caused by different physicochemical conditions in the gastro-
intestinal tract; although pH is neutral in D. magna gut, en-
zymatic conditions can be favorable to desorption processes.
However, this could not be the explanation for our results,
because the lower observed lethality impliesnonbioavailability
of the Mn in the pore-water particles.
To our knowledge, the concentrations of Mn found in the
sediments at several of the sampling stations are among the
highest for Mn reported in an environmental sample. The ques-
tion regarding the potential toxicity of such an amount of Mn
in the sediment is a complex one. We were concerned mainly
with elutriation phenomena that involved movement of Mn
from sediments to the water. Our results suggest that little of
the Mn from sediments is found in elutriates (pore water). In
the pore-water assessment of the Huitznopala and Tamala sta-
tions, the observed lethality was lower than expected (Table
4). This result can be explained by the extremely low bio-
availability of the Mn compounds in the particles within the
pore-water samples. However, these results need to be con-
sidered carefully, because we analyzed only 8 to 12 particles
for each of three sampling stations in three of the riversstudied.
A more complete analysis of particles from pore water might
be necessary to reinforce this argument.
An important goal of the present study was to evaluate the
toxicity of DDT, Mn, and coexposure to both toxicants. Clear-
ly, DDT was more toxic than Mn for D. magna, and the tox-
icant combination was more toxic than either DDT or Mn
alone. These results are quite relevant, because water-quality
criteria usually evaluate contaminants individually and not in
combination, as they often occur in the environment. For L.
quadridentata, the results were very different: The rotifer was
highly resistant to DDT and less susceptible to Mn compared
with D. magna. In the Mn/DDT coexposure, the dispersion of
the data gave us broad confidence limits that do not allow
conclusive results concerning the toxic behavior of the com-
Acknowledgement—The present work was supported by grant 1998-
0206013 of the Mexican Council for Science and Technology (CON-
ACYT) to F. Dı ´az-Barriga. The doctoral scholarship 83854 was grant-
ed to J. Mejı ´a-Saavedra by CONACYT.
1. Cassee RF, Groten PJ, Bladeren JP, Feron JV. 1998. Toxicological
evaluation and risk assessment of chemical mixtures. Crit Rev
2. Attar NE, Maly JE. 1982. Acute toxicity of cadmium, zinc, and
cadmium–zinc mixtures to Daphnia magna. Arch Environ Con-
tam Toxicol 11:291–296.
3. Biesinger KE, Christensen GM, Fiandt JT. 1986. Effects of metal
salt mixture on Daphnia magna reproduction. EcotoxicolEnviron
4. Steevens JA, Benson WH. 1997. Chemical mixture toxicity: Ef-
fects of chlorpyrifos, dieldrin, and methyl mercury on Hyalella
azteca. Abstracts, SETAC 18th Annual Meeting, San Francisco,
CA, USA, November 20, p 324.
5. Chang LW, Cockerham LG. 1994. Basic Environmental Toxi-
cology. CRC, Boca Raton, FL, USA.
6. Reynolds KD, Rainwater TR, Scollon EJ, Sathe SS, Adair BM,
Dixon KR, Cobb GP, McMurry ST. 2001. Accumulation of DDT
and mercury in prothonotary warblers ( Protonotaria citrea) for-
aging in a heterogeneously contaminated environment. Environ
Toxicol Chem 20:2903–2909.
7. Lewis MA, Daniels CB, Moore JC, Chen T. 2002. Potential gen-
otoxicity of wastewater-contaminated pore waters with compar-
ison to sediment toxicity and macrobenthic community compo-
sition. Environ Toxicol 17:63–73.
8. Eriksson PS. 2000. Temporal variations of manganese in the he-
molymph and tissues of the Norway lobster, Nephrops norvegicus
(L.). Aquat Toxicol 48:297–307.
9. World Health Organization. 1989. Environmental Health Criteria
83. DDT and Its Derivatives, Environmental Aspects. Geneva,
10. Hosokawa M, Endo G, Kuroda K. 1995. Acute toxic effect of
river Yodo water (Japan) on Daphnia magna. Bull Environ Con-
tam Toxicol 55:419–425.
11. Vigano ´ L. 2000. Assessment of the toxicity of River Po sediments
with Ceriodaphnia dubia. Aquat Toxicol 47:191–202.
12. Rico-Martı ´nez R, Vela ´zquez-Rojas CA, Pe ´rez-Legaspi IA, San-
tos-Medrano GE. 2000. The use of aquatic invertebrate toxicity
tests and invertebrate enzyme biomarkers to assess toxicity in the
states of Aguascalientes and Jalisco, Mexico. In Butterworth FM,
Gunatilake A, Gonsebatt ME, eds, Biomonitors and Biomarkers
as Indicators of Environmental Change, Vol 2. Plenum, New
York, NY, USA, pp 427–438.
13. Nikunen E, Miettinen V. 1985. Daphnia magna as an indicator
of the acute toxicity of waste waters. Bull Environ Contam Tox-
14. Zou E, Fingerman M. 1997. Synthetic estrogenic agents do not
interfere with sex differentiation but do inhibit molting of the
cladoceran Daphnia magna. Bull Environ Contam Toxicol 58:
15. Merrit CM, Dodson SI. 1998. An invertebrate ecologicalbioassay
for screening endocrine disruptors. Scientiae Naturae 1:11–25.
16. Baer KN, Owens KD. 1999. Evaluation of selected endocrine
disrupting compounds on sex determination in Daphnia magna
using reduced photoperiod and different feeding rates. Bull En-
viron Contam Toxicol 62:214–221.
17. Burbank S, Snell TW. 1994. Rapid toxicity assessment using
esterase biomarkers in Brachionus calyciflorus (Rotifera). En-
viron Toxicol Water Qual 9:171–178.
18. Juchelka CM, Snell TW. 1995. Rapid toxicity assessment using
ingestion rate of cladocerans and ciliates. Arch Environ Contam
19. U.S. Environmental Protection Agency. 1985. Methods for mea-
suring the acute toxicity of effluents to freshwater and marine
organisms. EPA-600/4-85-013. Washington, DC.
20. Pe ´rez-Legaspi IA, Rico-Martı ´nez R. 2001. Acute toxicity tests
on three species of the genus Lecane (Rotifera: Monogononta).
21. Ya ´nez L, Ortiz-Pe ´rez D, Batres LE, Borja-Aburto VH, Dı ´az-Bar-
riga F. 2002. Levels of dichlorodiphenyltrichloroethane and del-
tamethrin in humans and environmental samples in malarious
areas of Mexico. Environ Res Sect A 88:174–181.
22. Serrano L, Miracle MR, Serra M. 1986. Differential response of
Brachionus plicatillis (Rotifera) ecotypes to varios insecticides.
J Environ Biol 7:259–275.
23. Lotufo GR, Landrum PF, Gedeon ML, Tigue EA, Herche LR.
2000. Comparative toxicity and toxicokinetics of DDT and its
Environ. Toxicol. Chem. 24, 2005J. Mejı ´a-Saavedra et al.
major metabolites in freshwater amphipods. Environ Toxicol
24. Pillai MKK, Agarwal HC, Yadav DV. 1977. Tolerance, uptake,
and metabolism of DDT in Ganbusia affinis. Indian J Exp Biol
25. Ferrando MD, Andreu-Moliner E. 1991. Acute lethal toxicity of
some pesticides to Brachionus calyciflorus and Branchionus pli-
catilis. Bull Environ Contam Toxicol 47:479–484.
26. Ferna ´ndez-Casalderry A, Ferrando MD, Andreu-MolinerE.1992.
Acute toxicity of several pesticides to rotifer (Branchionus ca-
lyciflorus). Bull Environ Contam Toxicol 48:14–17.
27. Ferrando MD, Sancho E, Andreu-Moliner E. 1996. Chronic tox-
icity of fenitrothion to algae (Nannochloris oculata), a rotifer
(Branchionus calyciflorus), and the cladoceran(Daphniamagna).
Ecotoxicol Environ Saf 35:112–120.
28. Beauregard T, Ridal J. 2000. Evaluation of six simple bioassays
for the determination of drinking-water quality: Canadian results.
Environ Toxicol 15:304–311.
29. Rawash IA, Gaaboub IA, EL-Gayar EM, El-Shazli AY. 1975.
Standard curves for nuvacron, malathion, sevin, DDT, and kel-
thane tested against the mosquito Culex pipiens L. and the mi-
crocrustacean Daphnia magna Strauss. Toxicology 4:133–144.
30. Ziegenfuss PS, Renaudette WJ, Adams WJ. 1986. Methodology
for assessing the acute toxicity of chemicals sorbed to sediments:
Testing the equilibrium partitioning theory. In Poston TM, Purdy
R, eds, Aquatic Toxicology and Environmental Fate, Vol 9. STP
921. American Society for Testing and Materials, Philadelphia,
PA, pp 479–493.
31. Khangarot BS, Ray PK. 1989. Investigation of correlation be-
tween physicochemical properties of metals and their toxicity to
the water flea Daphnia magna Straus. Ecotoxicol Environ Saf
32. Erten-Unal M, Wixson GB, Gale N, Pitt LJ. 1998. Evaluation of
toxicity, bioavailability and speciation of lead, zinc and cadmium
in mine/mill wastewaters. Chem Speciat Bioavailab 10:37–46.
33. Weltens R, Goossens R, Van Puymbroeck S. 2000. Ecotoxicity
of contaminated suspended solids for filterfeeders(Daphniamag-
na). Arch Environ Contam Toxicol 39:315–323.
34. National Oceanic Atmospheric Administration. 1999. Screening
quick reference tables. Hazmat Report 99-1. Seattle, WA, USA.
35. Norma Oficial Mexicana. 2000. Modificacio ´n a la Norma Oficial
Mexicana NOM-127-SSA1-1994, Salud ambiental. Agua para
uso y consumo humano. Lı ´mites permisibles de calidad y trata-
mientos a que debe someterse el agua para su potabilizacio ´n.
Distrito Federal, Mexico.
36. Canadian Council of Ministers of the Environment. 2001. Ca-
nadian environmental water (sediment) quality guidelines for the
protection of aquatic life: 1999. Winnipeg, MB.