Bull Environ Contam Toxicol (2013) 91:191–196 DOI 10.1007/s00128-013-1029-0
Mercury Toxicity to Freshwater Organisms: Extrapolation Using Species Sensitivity Distribution Andreia C. M. Rodrigues • Fa´tima T. Jesus • Marco A. F. Fernandes • Fernando Morgado • Amadeu M. V. M. Soares • Sizenando N. Abreu
Received: 30 November 2012 / Accepted: 1 June 2013 / Published online: 16 June 2013 Ó Springer Science+Business Media New York 2013
Abstract Mercury toxicity to aquatic organisms was evaluated in different taxonomic groups showing the following species sensitivity gradient: Daphnia magna [ Daphnia longispina [ Pseudokirchneriella subcapitata [ Chlorella vulgaris [ Lemna minor [ Chironomus riparius. Toxicity values ranged from 3.49 lg/L (48 h-EC50 of D. magna) to 1.58 mg/L (48 h-EC50 of C. riparius). A species sensitivity distribution was used to estimate hazardous mercury concentration at 5 % level (HC5) and the predicted no effect concentration (PNEC). The HC5 was 3.18 lg Hg/L and the PNEC varied between 0.636 and 3.18 lg Hg/L, suggesting no risk of acute toxicity to algae, plants, crustaceans and insects in most freshwaters. Keywords Aquatic toxicity Mercury Species sensitivity distribution
Water pollution by mercury (Hg) is an issue of great concern due to its high toxicity, persistence in the environment, and potential for bioaccumulation through trophic chains. In addition, due to its high volatilization Hg can be transported to sites remotely located from point sources. In spite of imposed worldwide legislation urging to zero Hg discharges (OSPAR), anthropogenic Hg emissions in Asia, especially in China, are likely to increase significantly in the next decades (Lin et al. 2012). Therefore, water
A. C. M. Rodrigues (&) F. T. Jesus M. A. F. Fernandes F. Morgado A. M. V. M. Soares S. N. Abreu Department of Biology and CESAM, University of Aveiro, 3810-193 Aveiro, Portugal e-mail:
[email protected]
pollution by Hg is a priority and widespread environmental problem, which can impact both human and environmental health. Exposure of aquatic organisms to Hg can cause a multiplicity of adverse effects, arising from the binding of Hg ions to functional groups in proteins, namely sulfhydryl, phosphoryl, carboxyl, amide, and amine groups. Such interactions may cause protein precipitation, enzyme inhibition and corrosive action (Broussard et al. 2002). The majority of studies assessing Hg toxicity to aquatic organisms has focused on bioaccumulation and trophic transfer as well as lethal and sublethal toxicity to fish (Boening 2000); few studies addressed Hg toxicity to aquatic plants, and larval stages of insects (AzevedoPereira and Soares 2010; Dirilgen 2011). It is relevant to study Hg toxicity to a wide variety of organisms since they might be affected differently. Hg toxicity can be extrapolated to a broader range of species, using a species sensitivity distribution approach (SSD), which requires assembling of single-species toxicity. SSDs are one of the recommended approaches for ecological risk assessment and are used to predict hazardous concentrations (HC) affecting a certain percentage of species in a community. Commonly, this approach is used to determine HC5, the Hazard Concentration at 5 % level, i.e., the concentration that should protect 95 % of species. Following this approach, the predicted no effect concentration (PNEC) is also determined. This study aimed to assess Hg short-term toxicity to aquatic biota of several taxonomic groups, namely algae, plants, crustaceans and insects, and extrapolate to a broader range of aquatic species, using a SSD approach. Following this approach the HC5 and PNEC were determined. To the best of our knowledge, this is the first study addressing the ecological risk assessment of Hg based on SSDs.
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Materials and Methods Mercury was tested as mercury (II) chloride (HgCl2, Sigma-Aldrich, p. a. C99.0 %). A stock solution of HgCl2 was prepared with Milli-Q water and kept in the dark. Serial dilutions were prepared from the stock solution to the desired concentrations by diluting with the appropriate test medium for each species. The concentration of the stock solution and the tested metal concentrations at the beginning of exposures were certified with analysis by atomic absorption using the mercury analyzer AMA-254 (Altec, Czech Republic). The algae Pseudokirchneriella subcapitata and Chlorella vulgaris were cultured in unialgal batch cultures with sterilized MBL medium at 20 ± 1°C, under continuous and uniform cool-white light and continuous aeration. Growth inhibition tests for both species followed the OECD guideline 201 (OECD 2006a) and were followed in triplicate. Tests were initiated with 1.0 9 104 cells/mL in the log exponential growth phase. Tests were carried out in 250 mL Erlenmeyer flasks containing 200 mL of sterilized MBL medium with the desired Hg concentration. Test vials were randomly incubated in an orbital shaker (100 rpm) at 20 ± 1°C, with continuous light (cool-white fluorescent light, 4000 lux). After 72 h of exposure, the optical density at 440 nm (OD) was measured by spectrometry (Jenway 6505 UV/Visible spectrophotometer, UK) and converted to cell number employing the linear regression models previously developed for each species (P. subcapitata: Cell number = -171,075 ? OD 9 79,253,500; C. vulgaris: Cell number = -1,558,200 ? OD 9 131,443,240. Specific growth rates for each species were determined as the logarithmic increase in biomass, measured as cell number (OECD 2006a). Cultures of Lemna minor were maintained in Steinberg culture medium at 20 ± 1°C and photoperiod 16 h:8 h (light:dark) with light intensity about 6500 lux. Growth inhibition tests followed the OECD guideline 221(OECD 2006b), with three replicates per treatment. Three colonies with four visible fronds each were randomly assigned to each test vial (total 12 fronds per vial) containing 100 mL of Steinberg medium with the desired Hg concentration. Tests were carried out under the same conditions as cultures. Frond number was registered after 3 days (72 h) and 7 days of exposure. Dry weight was measured after 7 days of exposure by drying plants at 50°C during 24 h. Specific growth rates were determined considering both frond number and dry weight (OECD 2006b). Stock cultures of the crustaceans Daphnia magna [clone F sensu Baird et al. (1990)] and Daphnia longispina (clone EV20, sensu Antunes et al. (2003)] were maintained in ASTM hard water with a standard organic additive (Marinure seaweed extract, Glenside Organics Ltd.) and fed
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P. subcapitata (3.0 9 105 cells/mL for D. magna; 1.5 9 105 cells/mL for D. longispina). Temperature was 20 ± 1°C and photoperiod was 16 h:8 h (light:dark) both for cultures and tests. Acute immobilization toxicity tests with neonates (B24 h old) of D. magna and D. longispina followed the OECD guideline 202 (OECD 2004), with five replicates per treatment. Each replicate consisted of 5 organisms exposed to 50 mL of ASTM hard water with the desired Hg concentration and no food. The number of immobilized daphnids was recorded after 24 and 48 h of exposure. Cultures of Chironomus riparius larvae were kept in a 4L aquaria containing a layer of inorganic acid-washed fine sediment (B1 mm) as substrate and ASTM hard water, provided with aeration. Organisms were fed twice a week ad libitum with macerated fish flakes (Tetramin, Tetrawerke, Germany). Temperature was 20 ± 1°C and photoperiod was 16 h:8 h (light:dark) both for cultures and tests. Acute immobilization tests followed the OECD draft guideline for Chironomus sp. (OECD 2010), with 5 replicates per treatment. C. riparius larvae (B24 h old, instar I, 5 larvae per replicate) were exposed to Hg in Petri dishes containing 40 mL of ASTM with the desired Hg concentration and no food. The number of immobilized larvae was recorded after 24 and 48 h of exposure. To test if Hg caused a significant effect on the organisms compared with the appropriate control, a one-way ANOVA followed by a Dunnett test were pursued. Non-normally distributed or heteroscedastic data sets were analyzed with the nonparametric test Kruskal–Wallis followed by the Dunn’s post hoc test (SigmaPlot, v. 10, Systat Software Inc.). All statistical analyses were based on a 0.05 significance level. The 50 % effective concentrations (EC50) for continuous responses (algae and plant) were determined by adjusting data to binary responses (crustaceans and insects) were calculated with PriProbit software. For the SSD, toxicity data (EC50 values) were retrieved from the QSAR Toolbox (version 2.3, downloaded from http://www.qsartoolbox.org/). The experimental data of the present study was also considered. Multiple toxicity data for the same species were summarized as geometric means. Data was adjusted to a logprobit distribution and the HC5 determined. The SSD plot was generated using the EPA spreadsheet (SSD Generator V1, downloaded from http://www.epa.gov/caddis/da_ software_ssdmacro.html). The PNEC was calculated as the derived HC5 divided by a factor 1–5.
Results and Discussion The measured concentrations of freshly prepared test solutions were in good agreement with the nominal concentrations, as they did not differ more than 10 %.
Bull Environ Contam Toxicol (2013) 91:191–196 Fig. 1 Specific growth rate (day-1) of P. subcapitata and C. vulgaris after 72 h of exposure to Hg; Asterisk denotes a significant difference from the control
193
P. subcapitata
growth rate (day-1)
3.0
C. vulgaris
2.5
*
2.0
* *
*
1.5
* *
*
*
1.0
*
*
0.5 0.0 0
10
20
40
60
80
100
0
1
15
30
45
60
75
µg Hg /L
Hg effects on aquatic organisms followed dose–response curves (Figs. 1, 2, 3, 4). The EC50 values are summarized in Table 1. Cladocerans were the most sensitive group to Hg, and larvae of insects the least sensitive. Species sensitivity decreased in the following rank order: D. magna [ D. longispina [ P. subcapitata [ C. vulgaris [ L. minor [ C. riparius. The specific growth rate of both algae exposed to Hg is depicted in Fig. 1. P. subcapitata was twofold more sensitive to Hg than C. vulgaris (Table 1), which can be due to the development of resistance by C. vulgaris as suggested in a previous study (Juneau et al. 2001). The 72 h-EC50 toxicity value found for P. subcapitata compares well with the toxicity value 59 lg/L previously reported (USEPA 1985). Concerning C. vulgaris, the 72 h-EC50 value is lower than the value 400 lg/L reported by Rai et al. (1981) but the latter refers to a 3 weeks exposure period. Such a high toxicity value in the 3 weeks exposure is likely due to the development of resistance by this algae species, and agrees with the recovery of the photosynthetic activity of
0.20
growth rate (day-1)
frond number dry weight
0.15
*
0.10
*
* *
* *
0.05
*
*
0.00 0
0.4
0.8
1.2
1.6
2
2.4
mg Hg /L Fig. 2 Hg effects on the specific growth rate of L. minor measured as frond number (day-1) and dry weight (mg day-1) after 7 days of exposure; Asterisk denotes a significant difference from the control
C. vulgaris after 48 h of exposure to Hg, as reported by Juneau et al. (2001). Algae were the second most sensitive group to Hg. Such a high sensitivity has been reported as a result of photosynthetic inhibition (Juneau et al. 2001) and altered enzymatic activities (Jonsson and Aoyama 2009). The effects of Hg on the specific growth rate of L. minor, measured as both frond number and dry weight, are shown in Fig. 2. Hg effects were more pronounced during the first 3 days of exposure, as observed from the EC50 values in Table 1. During this period, plants exposed to 2.0 mg Hg/L and higher did not grow (considering frond number). The 7 days-EC50 value obtained in this study is in agreement with the value of 0.68 mg/L (95 % CI 0.536–0.881 mg/L) (Naumann et al. 2007), but is higher than the value 0.48 mg/L reported by Dirilgen (2011). The endpoint dry weight was more sensitive than frond number (Table 1), which is concordant with the reduced fronds’ size and chlorosis observed in plants exposed to high Hg concentrations. Indeed, plants grown in Hg concentrations 2 mg/L or higher exhibited small fronds; moreover, all plants grown in 2.4 mg Hg/L presented at least one white-yellowish frond, a sign of chlorosis. Immobilization of the crustaceans increased with increasing Hg concentration and exposure period, as shown in Fig. 3. D. magna was the most sensitive species to Hg, with a 48 h-EC50 value twofold lower than that of D. longispina. The 48 h-EC50 value for D. magna is in agreement with previous studies reporting values between 1.48 and 5.2 lg/L (USEPA 1985; Khangarot and Ray 1987). Concerning D. longispina, as far as we know, Hg toxicity to this species has not been determined previously. Such an absence of information is probably due to the fact this species is not very common for ecotoxicological studies, which contrasts with its wide geographical distribution (Seda et al. 2007). The immobilization of C. riparius larvae (I instar) after 24 and 48 h of exposure to Hg is depicted in Fig. 4. C. riparius were the least sensitive to Hg, with an EC50 value three orders of magnitude above those of crustaceans
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Bull Environ Contam Toxicol (2013) 91:191–196
Fig. 3 Immobilization of D. magna and D. longispina after 24 and 48 h of exposure to increasing Hg concentrations
D. magna
D. longispina
24h 48h
% immobilization
100 80 60 40 20 0 0
0.5
1
2
4
6
8
16
0
0.5
1
3
6
9
18
36
µg Hg /L
24h 48h
C. riparius
% immobilization
100 80 60 40 20 0 0
0.4
0.8
1.2
1.6
2
2.4
µg Hg /L Fig. 4 Immobilization of larvae of C. riparius after 24 and 48 h of exposure to increasing Hg concentrations
(Table 1). A previous study reported a 48 h-EC50 value of 3. 26 mg/L (Azevedo-Pereira and Soares 2010), but concerning IV instar larvae. As far as we know, Hg toxicity to this species at this stage has not been determined previously. The lower EC50 of instar I larvae compared to instar IV supports that larvae are more sensitive to Hg during the I instar (OECD 2010). In the aquatic environments larvae of C. riparius grow in the sediment, potentially exposed to higher Hg concentrations than in the water column, since Hg tends to adsorb to sediments (Lin et al. 2012). Indeed, concentrations as high as 0.3–2,100 mg Hg kg-1 dry weight were found in sediments of heavily contaminated sites in China (Lin et al. 2012). Thus, even though these organisms were the least sensitive to Hg, there might occur adverse effects if the organisms are exposed to highly contaminated sediments. The wide-ranging difference in Hg toxicity to both invertebrate groups suggests a specific mode of action and/ or different defense mechanisms. Note that organisms were exposed to Hg under the same conditions (medium, temperature, photoperiod, absence of food).
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The SSD plot is shown in Fig. 5. The predicted toxicity for the 5 % most sensitive organisms (HC5) is 3.18 lg Hg/L. In other words, Hg concentration up to 3.18 lg/L will protect 95 % of the species belonging to the studied groups. The derived PNEC (Predicted No Effect Concentration) varies between 0.636 and 3.18 lg Hg/L. This range is above Hg concentrations commonly found in the aquatic environment. Indeed, values between 0.01 and 12 ng Hg/L were reported in natural freshwaters around the world (Floyd et al. 2002). Thus, no acute toxicity is expected to algae, plants, crustaceans and larvae of insects, since these values are lower than the PNEC. However, in Chinese freshwater systems Hg concentrations as high as 0.69 lg/L were reported (Lin et al. 2012). Although the major part of PNEC range does not indicate risk of acute toxicity to aquatic organisms in these freshwaters, risk of acute toxicity will be expected when considering the lower limit of the PNEC (0.636 lg/L). In an acute pollution event, crustaceans (the most sensitive group to Hg) could be adversely affected, decreasing the feeding pressure on algae with potential consequences for ecosystem functioning by decoupling of trophic relationships (Domis et al. 2007). In the present study, acute toxicity data was used for Hg risk assessment. However, since the susceptibility of aquatic organisms to Hg is affected by water temperature and hardness (Boening 2000), we suggest that future contributions for Hg risk assessment should also include field experiments under varying temperature and hardness conditions. Moreover, chronic toxicity and bioaccumulation data should also be included in order to estimate risk for chronic exposures. This study is a contribution in the assessment of Hg effects in the aquatic systems. Organisms from different taxonomic groups showed different sensitivity to Hg. Cladocerans and larvae of insects were, respectively, the most sensitive and tolerant to Hg, with acute toxicity values differing in three orders of magnitude. Following the species sensitivity distribution, the PNEC varies between 0.636 and 3.18 lg Hg/L, which suggests that there is no
Bull Environ Contam Toxicol (2013) 91:191–196 Table 1 Hg toxicity values (EC50) to the test species
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Test species
Endpoint
P. subcapitata
GR
EC50 (95 % confidence interval) 24 h
48 h
72 h
7 days
–
–
51 lg/L
–
(37–66) C. vulgaris
GR
–
114 lg/L
–
–
(53–175) L. minor
FN
–
–
0.48 mg/L
0.81 mg/L
(0.40–0.56)
(0.63–0.99) 0.39 mg/L
DW
–
–
–
I
7.26 lg/L
3.49 lg/L
–
–
(4.96–9.56)
(2.41–4.57)
(0.24–0.55) D. magna
Fig. 5 SSD plot of the proportion of species affected versus mercury concentration. The filled symbols refer to experimental data of this study. HC5 = 0.00318 mg Hg/L (95 % CI 0.002–0.005)
I
10.76 lg/L (–)
7.14 lg/L (4.57–9.71)
–
–
C. riparius
I
2.29 mg/L
1.58 mg/L
–
–
(1.65–2.92)
(1.30–1.86)
1 centile 5% centile 95%
Insect
0.9
Proportion of species affected
GR growth rate, FN frond number, DW dry weight, I immobilization
D. longispina
Plantae
0.8
C. riparius L. minor
Crustacean Algae
0.7 0.6 0.5 0.4
C. vulgaris
0.3 0.2
D.longispina D. magna
P. subcapitata
r2 =0.982 ; n=32
0.1 0 0.00
0.01
0.10
1.00
10.00
100.00
mg Hg/L
risk of acute toxicity of Hg to algae, plants, crustaceans and larvae of insects in the majority of freshwaters at current environmental concentrations. Acknowledgments We thank Joa˜o Pedrosa, Taˆnia Vidal and Abel Ferreira (Department of Biology of the University of Aveiro) for providing C. riparius, L. minor cultures and algae cultures, respectively. The Portuguese Foundation for Science and Technology (FCT) supported the Post-doctoral fellowship SFRH/BPD/45807/2008 and CAPES- FCT 240/09 and the doctoral fellowship SFRH/BD/ 27637/2006.
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