SEARCH

SEARCH BY CITATION

Keywords:

  • acute-to-chronic ratio;
  • biofilm;
  • diuron;
  • ecotoxicology;
  • environmental risk assessment;
  • periphyton;
  • phytotoxicity;
  • species sensitivity distribution

Summary

  1. Top of page
  2. Summary
  3. Introduction
  4. Materials and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References
  • 1
    The assessment of the environmental hazards posed by chemical pollutants typically results from single-species tests that are extrapolated to ecosystems. The aim of the present study was to compare this type of extrapolation for a herbicide with the chronic effects that may be observed at a community level and to evaluate currently applied risk assessment strategies for their ability to predict chemical effects on complex communities.
  • 2
    Freshwater periphyton communities, grown in indoor aquaria, were exposed to the pollutant diuron for 3 months. Acute toxic effects of diuron were detected as photosynthesis inhibition using quenching analysis of chla-fluorescence. Chronic effects of the herbicide were observed in terms of changes in biomass and algal class composition as well as pollution-induced community tolerance (PICT). The PICT concept is based on a chemical exerting selection pressure on a community and therefore eliminating sensitive species. As a result, the measured community tolerance increases.
  • 3
    Short-term effects of diuron arise from 4–9 µg L−1 as half-maximal effect concentration (EC50). It is further shown that diuron concentrations down to 0·08 µg L−1 caused chronic effects in two independent microcosm studies. The observed threshold concentration of 0·08 µg L−1 still caused changes in biomass and class composition as well as an increased community tolerance. The determined EC50 values increased by a factor of 2–3 in diuron-exposed periphyton communities. This threshold value could not be predicted by advanced extrapolation methods such as species sensitivity distribution or acute-to-chronic effect ratios.
  • 4
     Synthesis and applications. The chronic community-level effects of the pollutant diuron were not predictable from single-species tests. However, regulations such as the EC Water Framework Directive or the EC-REACH process (Registrations, Evaluation and Authorisation of Chemicals) rely on this type of information. The management of chemicals in the environment should be based upon higher-tier assessment tools. Species interaction, detectable and quantifiable by the PICT methodology, may serve as a prognostic tool in chemical hazard assessment when extrapolating effects from single-species tests to community level.

Introduction

  1. Top of page
  2. Summary
  3. Introduction
  4. Materials and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References

A central purpose of environmental risk assessment is the protection of ecosystems from adverse impacts of chemicals. Because it is not possible to examine effects on ecosystems at large, single species, populations or communities are used to represent the ecosystem of interest. Micro- or mesocosms are used as model ecosystems under controlled conditions to simplify the studies of major ecosystem characteristics and observations are limited to a number of representative parameters. The observed effects are usually extrapolated to predict effects on natural ecosystems (Versteeg, Belanger & Carr 1999). Several approaches have been developed to reduce inherent uncertainty (Sijm et al. 2001). Predicted no-effect concentrations are commonly calculated using assessment factors (EC 2003).

Most of the data used in environmental risk assessment are based on the toxicological response at an individual or population level, including acute and chronic responses (Lin, Tokai & Nakanishi 2005), and are derived from single-species laboratory tests. Single-species tests are reproducible, but do not provide information on differences in species sensitivity or the effects of species interaction. The variation of sensitivities may be addressed by species sensitivity distributions (Kooijman 1987), which assume that variation in chemical sensitivity among species can be described by a statistical distribution (Roelofs et al. 2003). A species sensitivity distribution may be used to estimate the chemical concentration at which 95% of the species would be unaffected. This concentration is called 5% hazardous concentration (HC5) (Roelofs et al. 2003) and is used in probabilistic approaches to quantify environmental risks.

Further, effects on single species are usually recorded after exposure times of no more than one month, which is likely to be a shorter exposure time than that experienced by natural ecosystems. Therefore, chronic effects in ecosystems that might occur on a longer time scale are not addressed in these tests.

To assess variation between different exposure times, acute-to-chronic extrapolation, expressed as acute–chronic ratios (ACRs), have been proposed. The Technical Guidance Document (EC 2003) and ECETOC (European Centre for Ecotoxicology and Toxicology of Chemicals) aquatic toxicity data base define ACRs as ratios of an acute half-maximal effect concentration (EC50) vs. a chronic/subchronic no-effect concentration (NOEC) (Lange et al. 1998). Acute EC50 values usually refer to short-term exposure of invertebrates or fish for 48 h or 96 h, respectively. Chronic NOEC values usually originate from long-term exposures of invertebrates or fish for 21 days or 28 days, respectively.

Species interaction is an additional aspect of uncertainty when extrapolating from single-species tests to community- level effects. Contaminants may have the potential to change the outcome of species interactions and therefore influence community structure (Clements, Cherry & Cairns 1989), but species interaction is hardly ever considered in environmental risk assessment. One experimental approach to include effects caused by species interaction may be the use of community assays, representing one or more functional groups of an ecosystem. They possess a higher biological complexity with interacting species exposed simultaneously and all contributing to the observed response. Further, it is possible to monitor model ecosystems over prolonged time periods.

Species in a spatially structured environment, e.g. biofilms, are dependent on a complex web of symbiotic interactions (Hansen et al. 2007). They are characterized by competitive exclusion, predation or parasitism during succession but also by mutualism or commensalisms when coexisting in equilibrium. Additional stressors, e.g. toxicants may disturb this coexistence and induce a directed selection. Based on these considerations and according to the concept of pollution-induced community tolerance (PICT; Blanck & Wängberg 1988), it is possible to detect chronic effects from pollutants. The major mechanism behind PICT is the elimination of sensitive individuals or species and the proliferation of more tolerant ones during chronic exposure, meaning that the pollutant has exerted selection pressure on the sensitive members of the community (Nystrom, Bjornsater & Blanck 1999). Community tolerance can be examined by conducting several acute short-term tests. This offers the possibility to compare community responses to chronic exposure in terms of tolerance. Even more important, the observed effects include interactions within assemblages of various species, and therefore comprise several fundamental characteristics of natural communities that cannot be observed on a single-species level.

The scope of this present work was to compare commonly applied methods to calculate predicted no-effect concentrations in environmental risk assessment to effect-thresholds on community level. The objective was to account for species interaction as a factor of uncertainty in environmental risk assessment by applying the PICT concept. Periphyton was used to observe possible chronic effects over a period of several months. Tests were conducted using the phenylurea diuron, which provides a well-known mechanism of action and sufficient literature data on acute and chronic effects measured in single-species tests. It is a widely used herbicide, which acts as a photosystem II inhibitor (Devine, Duke & Fedtke 1993). This specific mechanism of action is universally effective in all autotrophic organisms and should therefore affect all species in periphyton communities; thus, differences in species sensitivity may derive from different uptake or transformation capabilities of the individual species. Environmental concentrations of diuron are usually below 0·1 µg L−1 but have exceeded 1 µg L−1 in surface near groundwater in Germany (LAWA 1997).

Materials and methods

  1. Top of page
  2. Summary
  3. Introduction
  4. Materials and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References

periphyton cultivation

Two consecutive long-term experiments over a period of 11 weeks and 12 weeks were conducted between January to June 2006. Periphyton, obtained from the Mulde river (contributor to the Elbe river, Germany), was grown on 1·5 cm2 glass discs that were arranged in plexiglass racks according to Blanck (1985). The light regime was a 14 h/10 h light/dark cycle using neon lamps (36 W/21–840; intensities above the water surface 200 µm photons m−2 s−1). The mean cultivation temperature was 21·6 ± 1·7 °C. The water was stirred continuously. No additional inoculation of photosynthetic microorganisms was added to the naturally obtained algae.

The water used for cultivation was taken unfiltered from the Mulde river, and to avoid nutrient limitation, water was replaced weekly. PH, oxygen and conductivity were measured once per week in fresh river water and in aquaria 1 day before water was exchanged. The experimental setup of both long-term experiments consisted of eight aquaria holding 20 L or 15 L river water with one replicate per treatment concentration and three replicates for untreated controls. Five aquaria were exposed to diuron (CAS RN 330-54-1, Lot 3268X, Riedel de Haen, Germany) in concentrations of 2, 0·4, 0·08, 0·016 and 0·0032 µg L−1 in the first experiment. Diuron exposure concentrations were chosen based on the effects shown in preliminary short-term photosynthesis-inhibition tests using periphyton. In the second experiment, five aquaria were exposed to diuron concentrations of 50, 10, 2, 0·4 and 0·08 µg L−1. For stock solution, diuron was dissolved in DMSO (CAS RN 67-68-5, Merck, Germany) and diluted with DMSO to the final concentrations. Three aquaria served as solvent controls with a DMSO concentration of 0·33% (v/v) in the first experiment and 0·5% (v/v) in the second experiment. These DMSO concentrations did not show effects in previous experiments. Diverse DMSO concentrations between the two experiments became necessary for technical reasons because of the different concentration ranges chosen for each experiment.

biomass and algal class determination

Biomass and algal class composition were analysed based on fluorescence measurements of photosynthesis pigments. Fluorescence was measured for three spots per sample with three samples taken from each of the aquaria. The basis of biomass determination with a pulse-amplitude modulated (PAM) fluorometer is the assumption that the minimal fluorescence yield F0 correlates well with the chlorophyll content and biomass (Jakob et al. 2005; Schmitt-Jansen & Altenburger 2008). Therefore, fluorescence measurements can be used to determine biomass of photosynthetic microorganisms.

There are several signature pigments for the different algal classes that allow for their discrimination by multi-wavelength excitation. For fluorescence measurements a Phyto-PAM fluorometer (Heinz Walz GmbH, Effeltrich, Germany) was used as previously described in Schmitt-Jansen & Altenburger (2008).

tolerance determination based on photosynthetic activity

The quantitative relationship between fluorescence and photosynthetic energy conversion is based on the fact that fluorescence originates from the same excited states, created by light absorption that can as well be photochemically converted or dissipated into heat. The measurement of variable chla-fluorescence does not only carry information on chla-content but also on the maximal and effective quantum yield of photosystem II, respectively (according to Schreiber, Schliwa & Bilger (1986)).

The effective quantum yield represents the photosynthetic capacity and can be used to assess the inhibition of photosystem II. A Maxi-Imaging-PAM-fluorometer (Heinz Walz GmbH, Effeltrich, Germany) was used for chla-fluorescence measurements. Chla-fluorescence was measured in three replicates, each by applying actinic light at an intensity set to 6 relative units for 2·5 min for sufficient adaptation to the actinic light and then applying a saturation pulse with the intensity set to 10. Chla-fluorescence, inline image and maximum fluorescence, inline image (according to Genty, Briantais & Baker (1989)) were measured just before and after the saturation pulse and were used to calculate the effective quantum yield.

Short-term tests were carried out with periphyton grown on glass discs several times during the long-term experiment. For this experiment, 24-well plates were used which allowed testing of six diuron concentrations, 100, 33, 11, 3·7, 1·2, and 0·4 µg L−1, with three replicates each in addition to six replicates for controls (three replicates containing DMSO and three replicates without DMSO). For the stock solution, 100 mg L−1 diuron was dissolved in DMSO. DMSO concentration was kept constant at 0·1% (v/v) in each well, except for controls without DMSO. Periphyton was exposed to diuron for 1 h on a rotary shaker at 200 r.p.m. with illumination of ~110 µm photons m−2 s−1 at 21 ± 2 °C. After incubation fluorescence of periphyton was measured with the MAXI-Imaging-PAM.

The determined inline image and inline image are suitable to calculate the effective quantum yield as

  • image

The quantum yield is used to calculate the inhibition of photosystem II activity of exposed samples in comparison to controls using the following formula:

  • image

statistics

Photosynthesis inhibition data measured in short-term tests were fitted assuming logistic distribution of the data using the equation

  • image

where E equals the effect, Amin and Amax denote the minimal and maximal effect, and p stands for the slope. The obtained data allowed Amin and Amax to be set to 0 and 100, respectively. EC50 is the calculated toxicant concentration causing 50% inhibition of the test parameter. All statistical parameters were estimated using the software Origin 7·5 (OriginLab, Northampton, MA, USA). The lowest observed effect concentration (LOEC) was determined as the lowest tested concentration of the chemical that produced a distinguishable effect on periphyton, in comparison to baseline toxicity. The no-effect concentration was determined as the highest applied concentration of the chemical that did not produce any measurable effect on periphyton.

Results

  1. Top of page
  2. Summary
  3. Introduction
  4. Materials and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References

The pH value increased from minimally 7·1 in fresh river water to maximally 10·9 in aquaria at the end of 1 week. The mean pH value in fresh river water was 7·7 (SD 0·9) and in aquaria after 1 week 9·5 (SD 0·8). The increase in pH is due to the photosynthetic conversion of carbon dioxide reducing the carbonate content and therefore increasing pH. This phenomenon is well described in aquatic systems as well as for microbial mats (Revsbech et al. 1983). The microcosm experiment aimed to reconstruct standardized but semi-natural conditions; therefore, pH was not buffered and natural fluctuations of pH were accepted. Oxygen content ranged from 6·0–18·8 mg L−1 with a mean value of 9·1 mg L−1 (SD 1·9 mg L−1). Conductibility ranged from 269–757 µS with a mean conductibility of 466 µS (SD 124 µS).

To determine acute effects of diuron on photosynthesis, 20 short-term tests were conducted using periphyton cultivated without addition of diuron at different times during the long-term experiments of 3 to 12 weeks. As expected from its mechanism of action, diuron caused inhibition of photosynthesis at low concentrations in the µg L−1 range. EC50 ranged between 2·6 and 15·2 µg L−1 diuron with 80% of the values ranging between 4 and 9 µg L−1 diuron. According to Blanck (2002), this is termed the baseline tolerance of unexposed periphyton against diuron and will be used as reference for further results. The EC50 of 15·2 µg L−1 marks an outlier. The slope values range from 1·4 to 3·0 with a median value of 1·9. The determined LOEC was 1·2 µg L−1, the NOEC was 0·4 µg L−1.

chronic effects

Chronic effects of diuron were determined after long-term exposure over a period of several weeks. During the first experiment (I), several short-term inhibition tests were conducted at three different times to monitor tolerance development of periphyton. The EC50 values of controls were within the range of the baseline toxicity at all times and did not show any significant correlation with age or minimum fluorescence of the periphyton (data not shown). After 4 weeks, periphyton grown with diuron concentrations between 0–0·4 µg L−1 denoted no increases in tolerance compared to the baseline tolerance (Table 1). Diuron pre-exposure at 2 µg L−1 increased tolerance after 4 weeks up to an EC50 of 23·6 µg L−1. After 7 weeks of growth, an increased tolerance was determined for diuron concentrations between 0·08–2 µg L−1. The determined EC50 values show a strong correlation with the pre-exposure concentration of diuron (Fig. 1). After 11 weeks of growth, an increased tolerance was no longer determined for any diuron concentrations. These results show a minimal concentration of diuron to increase periphyton tolerance at 0·08 µg L−1 diuron during cultivation.

Table 1.  Overview of EC50-values (µg L-1) in both experiments for all diuron concentrations applied during cultivation (n.a. not available; n.d. not determinable); enhanced EC50-values in comparison to baseline toxicity are highlighted; baseline toxicity (4-9 µg L-1) was calculated from short-term tests from control aquaria at different times throughout the experiment.
Chronic diuron concentration (µg L-1)EC50 values (µg L-1)
Experiment IExperiment II
Controls (3-12 weeks)4-9
 after 4 weeks of growthafter 3 weeks of growth
0.00322.0 ± 0.3n.a.
0.0164.9 ± 0.4n.a.
0.08n.d.4.4 ± 0.7
0.45.0 ± 1.33.2 ± 0.5
223.6 ± 6.34.7 ± 1.8
10n.a.3.0 ± 0.9
50n.a.n.d.
 after 7 weeks of growth 
0.00326.9 ± 0.6 
0.0168.3 ± 0.8 
0.0812.5 ± 1.1 
0.412.6 ± 0.5 
216.9 ± 3.5 
10n.a. 
50n.a. 
 after 11 weeks of growthafter 12 weeks of growth
0.00325.4 ± 0.4n.a.
0.0167.1 ± 0.6n.a.
0.084.7 ± 0.721.4 ± 0.7
0.45.1 ± 1.416.6 ± 0.5
25.8 ± 0.822.8 ± 0.7
10n.a.26.0 ± 3.6
50n.a.n.d.
image

Figure 1. EC50 concentrations of diuron after 1 h incubation as calculated from concentration-response curves of photosynthetic activity from 7-week-old periphyton, pre-exposed to diuron and solvent controls. (inline image Diuron pre-exposed periphyton – logarithmic fit y = ln(x − 0·4) R2 = 0·941).

Download figure to PowerPoint

The highest possible diuron concentration that would increase tolerance was assessed in the second experiment (II). After 3 weeks of growth, the EC50 values were determined for all concentrations, but no increase in tolerance was measured compared to baseline tolerance at this stage of community development. After 12 weeks of exposure, the tolerance increased for concentrations between 0·08–10 µg L−1 diuron up to EC50 values of 26 µg L−1 (Fig. 2).

image

Figure 2. Concentration-response curves of photosynthetic activity against diuron after 1 h incubation as measured with periphyton that was pre-exposed to diuron for 12 weeks and solvent controls from experiment II (curves 1 and 2: solvent controls; curve 3: 0·4 µg L−1; curve 4: 0·08 µg L−1; curve 5: 2 µg L−1; curve 6: 10 µg L−1 diuron).

Download figure to PowerPoint

For periphyton cultivated with 50 µg L−1 diuron, no tolerance could be determined at any time, as the data were not suitable for a logistic fit. Periphyton was severely disturbed; thus, a concentration of 50 µg L−1 diuron was not tolerable to periphyton. As shown in Table 1, diuron concentrations between 0·08–2 µg L−1 increased tolerance in both experiments.

In the second experiment, chronic effects of diuron were also monitored in terms of the development of chla-fluorescence and algal class composition using a multi-wavelength excitation PAM. Chla-fluorescence was similar in diuron-exposed and non-exposed aquaria until 5 weeks of growth. Thereafter, periphyton exposed to 0·08–2 µg L−1 diuron started to increase in minimum fluorescence much more than non-exposed periphyton for a further 2 weeks and then reached a steady state. The non-exposed periphyton started to increase in chla-fluorescence after 5 weeks, but reached a lower fluorescence level (examples shown in Fig. 3). Fluorescence of periphyton exposed to 10 µg L−1 diuron did not differ from non-exposed periphyton (Fig. 4). Exposure to 50 µg L−1 diuron caused a slight inhibition of periphyton fluorescence compared to non-exposed periphyton. Fluorescence patterns after 12 weeks for the whole dilution series is shown in Fig. 4.

image

Figure 3. Periphyton minimum fluorescence (F0) of Chla over time for two representative examples, measured with a multi-wavelength-excitation PAM-fluorometer (inline image periphyton with 0·4 µg L−1 diuron pre-exposure; inline image non-exposed periphyton).

Download figure to PowerPoint

image

Figure 4. Total periphyton minimum fluorescence (F0) of chla for different pre-exposure concentrations of diuron, measured with a multi-wavelength-excitation PAM-fluorometer.

Download figure to PowerPoint

The algal class composition of periphyton was monitored over the whole period of 12 weeks. The non-exposed periphyton consisted of about 90% diatoms and 10% cyanobacteria (Fig. 5). The periphyton grown with low diuron concentrations between 0·08–2 µg L−1 started to differ from non-exposed periphyton after 8 weeks. The relative amount of diatoms decreased to about 50%, whereas the amount of cyanobacteria increased to about 40%. Only 0·08 µg L−1 diuron caused a change in the relative amount of green algae, which increased to about 20%. The high diuron concentrations of 10 and 50 µg L−1 did not cause clear changes in the algal class composition of periphyton biomass.

image

Figure 5. Algal class composition of total biomass over time (inline image, control 1; inline image, control 3; inline image, 0·08 µg L−1 diuron; inline image, 0·4 µg L−1 diuron; inline image, 2 µg L−1 diuron).

Download figure to PowerPoint

Discussion

  1. Top of page
  2. Summary
  3. Introduction
  4. Materials and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References

In the present study, acute and chronic effects of diuron on freshwater periphyton over a period of 3 months were observed in terms of photosynthesis inhibition, chla-fluorescence, community structure and tolerance development (PICT). Acute effects of diuron on photosynthesis inhibition, measured as variable chla-fluorescence in short-term tests, showed maximal effect concentration (EC50 values) of 4–9 µg L−1. This range included 80% of all the EC50 values determined in 20 independent short-term tests. EC50 concentrations for photosynthesis inhibition of marine microalgae communities, measured as incorporation of 14C, ranged between 4–20 µg L−1 diuron (Molander & Blanck 1992; Blanck & Dahl 1996; Arrhenius et al. 2004; Backhaus, Arrhenius & Blanck 2004) and 20–51 µg L−1 for stream periphyton (Dorigo et al. 2007). This consistency of the EC50 concentrations is in agreement with the specific mechanism of action of diuron and the high conservation of photosynthetic reaction centres within the plant kingdom (Bengtson Nash et al. 2005).

The observed chronic effects of diuron on periphyton included a clear increase of community tolerance for concentrations in long-term experiments between 0·08–10 µg L−1. Further, changes in algal class composition and increase of chla-fluorescence were observed for concentrations between 0·08–2 µg L−1. Minimum fluorescence was used as a proxy of chla in this study. An increase of the chla-content of algal cells after exposure to photosystem II inhibitors is a well-known phenomenon. Whether the observed increase in chla of exposed periphyton derived from a relative increase of chla per cell or of enhanced growth of algal cells remains unclear. From optical inspection, a higher density of periphyton cells seems plausible. The applied diuron concentration of 50 µg L−1 marks a highly toxic concentration that inhibited growth by almost 100%. The wide concentration range for the chronic exposure studies (0·0032–50 µg L−1) was chosen to cover the boundaries of induced tolerance. Environmental concentrations of diuron in surface near groundwater, reported by German water authorities (LAWA 1997) exceeded the PICT thresholds of this study, indicating that community shifts may occur in the environment.

However, differences between the two experiments were found regarding the time dependence of the observed PICT effect. The results originate from two independent but comparable long-term studies, which cultivated periphyton from a natural water source. Communities grown in natural water are subject to changes in water quality (e.g. nutrients) during their development. Therefore, the abundances and activities of different species, that might have different sensitivities, may change over time. As a result, in different studies the same point in time from the start of the experiment will probably resemble a different developmental stage of the community, depending on the composition of the inoculation. Additionally, detachment of large pieces of biofilm from the surface (sloughing) may be a confounding factor of biofilm succession, which relegates the biofilm to an initial succession phase. This will lead to varying EC50 values as found at similar times during the two experiments. Despite these variations, the effects caused by defined concentrations of diuron were highly reproducible, in terms of a certain concentration causing PICT or changes in biomass composition and abundance at any point in time during the study. Changes in composition of algal classes occur for the same diuron concentrations as the increase in community tolerance, clearly supporting the concept that more tolerant organisms will replace sensitive ones when a toxicant exerts selection pressure on the community. This was shown at a species level (Schmitt-Jansen & Altenburger 2005) and was detectable as a shift of algal classes.

Comparable studies were conducted by Molander & Blanck (1992) analysing PICT of marine periphyton for diuron and in combination with tri-n-butyl tin (TBT) (Molander et al. 1992). The threshold concentration causing PICT in both studies was around 10 µg L−1 diuron, which ranges about two orders of magnitude higher than the results of the present study. The process of replacement of sensitive species by less sensitive species due to selection pressure may take several reproduction cycles before PICT can be observed, despite the comparatively short generation time of unicellular algae and the fast action of diuron. Both studies by Molander & Blanck (1992) and Molander et al. (1992) showed that an observable PICT effect can be expected after 3–4 weeks when using a flow-through system, which allows new species to mix with the existing community at all times. However, the PICT tests of the present study conducted after 3 or 4 weeks, respectively, revealed no induced tolerance after 3 weeks, and after 4 weeks tolerance was only increased for 2 µg L−1 diuron pre-exposure. This may be due to the semi-static design of the experiments, leading to longer succession times than a flow-through system. In addition, it may be argued that 4 weeks might not be long enough to allow for the induction of tolerance by lower concentrations of diuron, as was found in this study after 7 weeks and 12 weeks of exposure, respectively. This means that diuron concentrations below the acute LOEC of 1·2 µg L−1 may need a longer period of exposure before PICT effects could be observed. Low PICT thresholds in the field were recently found by Dorigo et al. (2007), with a clear correlation of EC50 values for field periphyton and total pesticide concentration. Environmental concentrations of diuron ranged between 0·09 and 0·43 µg L−1 in that study, which was in the range the present study could detect PICT in microcosms.

Several other PICT studies have been conducted for heavy metals such as zinc, copper or arsenic, and triazines, such as atrazine and Irgarol 1051, or TBT (Table 2). For most of these chemicals, the EC50 concentrations in acute tests are about one to two orders of magnitude higher than the concentrations found to induce PICT in periphyton. In accordance with these, the present study revealed that for diuron, PICT appeared distinctly more sensitive than acute tests, in that the PICT threshold concentration was two orders of magnitude below the observed acute EC50 concentrations.

Table 2.  Overview of available data for chemical substances tested in a PICT study; observed NOECs originate from algal growth tests; ACRs were calculated as ratios between EC50 and NOEC originating from the same study; acute EC50 originates from algae photosynthesis inhibition after ≤ 1 h; predicted chronic NOECs were calculated from acute EC50 applying the ACR from growth test; asterisk (*) indicates data based on marine communities
Chemical substanceAcute EC50 for algae single-species (photosynthesis) (µg L−1) (reference)Calculated ACR from algae growth test (reference)Predicted chronic NOEC (µg L−1)Observed NOEC from algae growth test (µg L−1) (reference)Lowest chronic concentration that induced PICT (µg L−1) (reference)
Atrazine141 (Altenburger et al. 1990)5 (Faust et al. 2001)288 (Faust et al. 2001)10 (Berard & Benninghoff 2001)
Irgarol 10511 (Arrhenius et al. 2006)11 (Arrhenius et al. 2006)0·090·5 (Arrhenius et al. 2006)0·13 (Nystrom et al. 2002)
TBT17 (Arrhenius et al. 2006)1·6 (Arrhenius et al. 2006)939 (Arrhenius et al. 2006)0·14 (Blanck & Dahl 1998)*
Copper1968 (van der Heever & Grobbelaar 1996)2·1 (Schafer et al. 1994)93756 (Schafer et al. 1994)0·94 (Gustavson & Wängberg 1995)
Diuron11 (El Jay et al. 1997)9 (Backhaus et al. 2004)1·20·7 (Backhaus et al. 2004)0·08 (present study)

Because it is very cost- and labour-intensive to conduct community-level studies for every potentially hazardous chemical, extrapolation methods of chemical effects on communities from single-species results are needed for environmental risk assessments. They should consider specific community characteristics (e.g. variation in species sensitivity), which can differ among algal species by up to several orders of magnitude (Blanck 1984; Nystrom et al. 1999). Acute EC50 concentrations of diuron regarding photosynthesis inhibition range from 2–233 µg L−1 between different algal species (Holliste & Walsh 1973; El Jay et al. 1997; Schreiber et al. 2002; Bengtson Nash et al. 2005; Podola & Melkonian 2005). A popular method in environmental risk assessment is to estimate 5% hazardous concentration of chemicals from species sensitivity distribution. For diuron, the HC5 based on a log logistic distribution of EC50 concentrations regarding acute photosynthesis inhibition is 3·4 µg L−1 (Fig. 6). This value is in line with the EC50 values found for short-term exposure of communities in this study. The HC5 value based on a log logistic distribution of NOECs is 0·74 µg L−1, which ranges between the acute LOEC and NOEC found in the present study. At the same time, both these concentrations are distinctly higher than the threshold concentration of 0·08 µg L−1 diuron that induced tolerance. Consequently, the evaluation of species sensitivity variation using a statistic species sensitivity distribution may represent different species sensitivities, but does not sufficiently illustrate the effects observed in communities.

image

Figure 6. Species sensitivity distributions for diuron; data from single species algae assays were derived from the RIVM e-toxBase database (de Zwart, personal communication); included were only EC50, respectively NOEC, concentrations regarding acute photosynthesis inhibition.

Download figure to PowerPoint

Another factor that may be considered for extrapolation is variation in exposure time, such as acute-to-chronic ratios. A single-species test with diuron was conducted by Backhaus et al. (2004) using a unicellular green algae to assess the inhibition of reproduction after 24 h. The ACR calculated as the ratio of EC50 to NOEC was 9·4. Several data from other single-species assays also show ACR values around 10 (Isnard 1998; Roelofs et al. 2003; Ahlers et al. 2006). This consistency of ACRs indicates that in single-species assays, acute effects appear about one order of magnitude above the concentrations that typically cause chronic effects on single species. If effects from single species are to be extrapolated to communities, the question is whether these ACRs sufficiently describe chronic effects on the community level. This was tested for several photosystem II inhibitors (Table 2).

The calculated ACRs were applied to acute EC50 values (1–3 h photosynthesis inhibition) for algae, which resulted in the listed predicted chronic NOEC. In addition, the observed NOEC values are listed. All data are based on freshwater algae, except one PICT study for TBT. In this data set, the comparison of freshwater single-species tests and marine communities may confound the findings; however, with regard to the principal differences in test conditions between community and single-species studies, this factor may be of minor relevance. For atrazine and Irgarol 1051 the predicted and observed NOEC are in the same order of magnitude as concentrations that caused chronic effects on community level (PICT). In contrast, for TBT, copper and diuron the predicted and observed NOECs are at least one order of magnitude above concentrations that caused PICT. Therefore, an extrapolation from acute single-species effects to chronic effects on community level using acute-chronic ratios has its limitations.

In conclusion, neither variation in species sensitivities nor differences in exposure time could adequately predict the low threshold concentration of diuron that caused chronic effects on the community level. One factor that is not covered by species sensitivity variation or acute-to-chronic extrapolation is the interaction between species, which is an essential organizing force in communities. It is reported from a variety of autotrophic and heterotrophic biofilms that species interaction changes the coexistence of component species, community structure and function by resource competition or chemical interaction (e.g. allelopathic activity; Van der Grinten 2005; Hansen et al. 2007); thus, community tolerance can be regarded as a quantifiable surrogate of species interaction.

management implications

The results of the present study show that species interaction might account for differences between effect concentrations observed for single species vs. communities. It is shown that chronic community-level effects of diuron were not predictable from single-species tests. However, regulations such as the EC water framework directive (WFD) or the EC-REACH process rely on this type of information. Within the EC-WFD, Environmental Quality Standard (EQS) were defined for a set of priority pollutants and are now applied all over Europe. For diuron, an EQS of 0·2 µg L−1 is defined (EQS data sheet 13, Brussels 2005); however, this study has shown that this will not protect microalgal communities, indicating that water managers should consider higher-tier studies for the formulation of EQSs. Model ecosystems, e.g. micro- or mesocosms including several species of different sensitivity that interact with each other, but also PICT field studies (Dorigo et al. 2007) may provide a useful and complementary tool to current definition of EQSs. Alternatively, an additional safety factor could be used to account for species interaction effects at the community level and reduced if appropriate experimentations were done. Literature reviews revealed similar findings for other chemicals, indicating that species interaction might be a sensitive parameter for adverse effects on communities. Further PICT studies as a surrogate of species interaction should allow these findings to be applied and generalized to other chemicals. They may provide a prognostic tool for risk assessment of chemicals in the environment to reduce uncertainty when intending to extrapolate chemical effects from single-species tests to a community level in chemical hazard assessment.

Acknowledgements

  1. Top of page
  2. Summary
  3. Introduction
  4. Materials and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References

We would like to thank D. de Zwart (RIVM, National Institute for Public Health and the Environment, The Netherlands) for providing algal effect data of diuron and T. Backhaus and an anonymous reviewer for their helpful comments on the manuscript. Thanks also go to J. Krüger, I. Christmann and S. Aulhorn for expert technical assistance. This study was funded by the Commission of the European Community (MODELKEY, 511237-GOCE) and (KEYBIOEFFECTS, MRTN-CT-2006-035695).

References

  1. Top of page
  2. Summary
  3. Introduction
  4. Materials and methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References
  • Ahlers, J., Riedhammer, C., Vogliano, M., Ebert, R.U., Kuhne, R. & Schuurmann, G. (2006) Acute to chronic ratios in aquatic toxicity – variation across trophic levels and relationship with chemical structure. Environmental Toxicology and Chemistry, 25, 29372945.
  • Altenburger, R., Bödeker, W., Faust, M. & Grimme, L.H. (1990) Evaluation of the isobologram method for the assessment of mixtures of chemicals. Ecotoxicology and Environmental Safety, 20, 98114.
  • Arrhenius, A., Backhaus, T., Gronvall, F., Junghans, M., Scholze, M. & Blanck, H. (2006) Effects of three antifouling agents on algal communities and algal reproduction: mixture toxicity studies with TBT, Irgarol, and Sea-Nine. Archives of Environmental Contamination and Toxicology, 50, 335345.
  • Arrhenius, A., Gronvall, F., Scholze, M., Backhaus, T. & Blanck, H. (2004) Predictability of the mixture toxicity of 12 similarly acting congeneric inhibitors of photosystem II in marine periphyton and epipsammon communities. Aquatic Toxicology, 68, 351367.
  • Backhaus, T., Arrhenius, A. & Blanck, H. (2004) Toxicity of a mixture of dissimilarly acting substances to natural algal communities: predictive power and limitations of independent action and concentration addition. Environmental Science & Technology, 38, 63636370.
  • Backhaus, T., Faust, M., Scholze, M., Gramatica, P., Vighi, M. & Grimme, L.H. (2004) Joint algal toxicity of phenylurea herbicides is equally predictable by concentration addition and independent action. Environmental Toxicology and Chemistry, 23, 258264.
  • Bengtson Nash, S.M., Quayle, P.A., Schreiber, U. & Müller, F. (2005) The selection of a model microalgal species as biomaterial for a novel aquatic phytotoxicity assay. Aquatic Toxicology, 72, 315326.
  • Berard, A. & Benninghoff, C. (2001) Pollution-induced community tolerance and seasonal variations in the sensitivity of phytoplankton to atrazine in nanocosms. Chemosphere, 45, 427437.
  • Blanck, H. (1984) Species-dependent variation among aquatic organisms in their sensitivity to chemicals. Ecological Bulletins, 36, 107119.
  • Blanck, H. (1985) A simple, community level, ecotoxicological test system using samples of periphyton. Hydrobiologia, 124, 251261.
  • Blanck, H. (2002) A critical review of procedures and approaches used for assessing pollution-induced community tolerance (PICT) in biotic communities. Human and Ecological Risk Assessment 8, 10031034.
  • Blanck, H. & Dahl, B. (1996) Pollution-induced community tolerance (PICT) in marine periphyton in a gradient of tri-n-butyltin (TBT) contamination. Aquatic Toxicology, 35, 5977.
  • Blanck, H. & Dahl, B. (1998) Recovery of marine periphyton communities around a Swedish marina after the ban of TBT use in antifouling paint. Marine Pollution Bulletin, 36, 437442.
  • Blanck, H. & Wängberg, S.-A. (1988) Pollution-induced community tolerance. A new ecotoxicological tool. In: Functional Testing of Aquatic Biota for Estimating Hazards of Chemicals (eds J.Cairns & J.R.Pratt), pp. 219230. American Society for Testing Materials, Philadelphia, PA, USA.
  • Clements, W.H., Cherry, D.S. & Cairns, J. Jr (1989) The influence of copper exposure on predator–prey interactions in aquatic insect communities. Freshwater Biology, 21 (3), 483488.
  • Devine, M., Duke, S.O. & Fedtke, C. (1993) Physiology of Herbicide Action. Prentice-Hall, Englewood Cliffs, NJ, USA.
  • Dorigo, U., Leboulanger, C., Berard, A., Bouchez, A., Humbert, J.F., Montuelle, B. (2007) Lotic biofilm community structure and pesticide tolerance along a contamination gradient in a vineyard area. Aquatic Microbial Ecology, 50, 91102.
  • EC (2003) Environmental Risk Assessment. Technical guidance document in support of Commission Directive 93/67/EEC on risk assessment for new notified substances and Commission Regulation (EC) no. 1488/94 on risk assessment for existing substances. European Commission, Brussels, Belgium.
  • El Jay, A., Ducruet, J-M., Duval, J-C. & Pelletier, J.P. (1997) A high sensitivity chlorophyll fluorescence assay for monitoring herbicide inhibition of photosystem II in the chlorophyte Selenastrum capricornutum: comparison with effect on cell growth. Archives of Hydrobiology, 140, 273286.
  • Faust, M., Altenburger, R., Backhaus, T., Blanck, H., Boedeker, W., Gramatica, P., Hamer, V., Scholze, M., Vighi, M. & Grimme, L.H. (2001) Predicting the joint algal toxicity of multi-component s-triazine mixtures at low-effect concentrations of individual toxicants. Aquatic Toxicology, 56, 1332.
  • Genty, B., Briantais, J.-M. & Baker, N.R. (1989) The relationship between the quantum yield of photosynthetic electron transport and quenching of chlorophyll fluorescence. Biochimica et Biophysica Acta, 990, 8792.
  • Gustavson, K. & Wängberg, S.-A. (1995) Tolerance induction and succession in microalgae communities exposed to copper and atrazine. Aquatic Toxicology, 32, 283302.
  • Hansen, S.K., Rainey, P.B., Haagensen, J.A.J. & Molin, S. (2007) Evolution of species interaction in a phototrophic biofilm community. Nature, 445, 533536.
  • Holliste, T.A. & Walsh, G.E. (1973) Differential responses of marine phytoplankton to herbicides – oxygen evolution. Bulletin of Environmental Contamination and Toxicology, 9, 291295.
  • Isnard, P. (1998) Assessing the environmental impact of wastewaters. Ecotoxicology and Environmental Safety, 40, 8893.
  • Jakob, T., Schreiber, U., Kirchesch, V. & Langer, U. (2005) Estimation of chlorophyll content and daily primary production of the major algal groups by means of multiwavelength-excitation PAM chlorophyll fluorometry: performance and methodological limits. Photosynthesis Research, 83, 343361.
  • Kooijman, S.A.L.M. (1987) A safety factor for Lc50 values allowing for differences in sensitivity among species. Water Research, 21, 269276.
  • Lange, R., Hutchinson, T.H., Scholz, N. & Solbe, J. (1998) Analysis of the ECETOC aquatic toxicity (EAT) database. II. Comparison of acute to chronic ratios for various aquatic organisms and chemical substances. Chemosphere, 36, 115127.
  • LAWA (1997) Zielvorgaben zum Schutz oberirdischer Binnengewässer, Volume 1, Laenderarbeitsgemeinschaft Wasser, Berlin.
  • Lin, B.L., Tokai, A. & Nakanishi, J. (2005) Approaches for establishing predicted-no-effect concentrations for population-level ecological risk assessment in the context of chemical substances management. Environmental Science & Technology, 39, 48334840.
  • Molander, S. & Blanck, H. (1992) Detection of pollution-induced community tolerance (PICT) in marine periphyton communities established under Diuron exposure. Aquatic Toxicology, 22, 129144.
  • Molander, S., Dahl, B., Blanck, H., Jonsson, J. & Sjostrom, M. (1992) Combined effects of Tri-Normal-Butyl Tin (Tbt) and diuron on marine periphyton communities detected as pollution-induced community tolerance. Archives of Environmental Contamination and Toxicology, 22, 419427.
  • Nystrom, B., Becker-van Slooten, K., Berard, A., Grandjean, D., Druart, J.C. & Leboulanger, C. (2002) Toxic effects of Irgarol 1051 on phytoplankton and macrophytes in Lake Geneva. Water Research, 36, 20202028.
  • Nystrom, B., Bjornsater, B. & Blanck, H. (1999) Effects of sulfonylurea herbicides on non-target aquatic micro-organisms – growth inhibition of micro-algae and short-term inhibition of adenine and thymidine incorporation in periphyton communities. Aquatic Toxicology, 47, 922.
  • Podola, B. & Melkonian, M. (2005) Selective real-time herbicide monitoring by an array chip biosensor employing diverse microalgae. Journal of Applied Phycology, 17, 261271.
  • Revsbech, N.P., Jorgensen, B.B., Blackburn, T.H. & Cohen, Y. (1983) Microelectrode studies of the photosynthesis and O2, H2S and pH profiles of a microbial mat. Limnological Oceanography, 28 (6), 10621074.
  • Roelofs, W., Huijbregts, M.A.J., Jager, T. & Ragas, A.M.J. (2003) Prediction of ecological no-effect concentrations for initial risk assessment: combining substance-specific data and database information. Environmental Toxicology and Chemistry, 22, 13871393.
  • Schafer, H., Hettler, H., Fritsche, U., Pitzen, G., Roderer, G. & Wenzel, A. (1994) Biotests using unicellular algae and ciliates for predicting long-term effects of toxicants. Ecotoxicology and Environmental Safety, 27, 6481.
  • Schmitt-Jansen, M. & Altenburger, R. (2005) Toxic effects of isoproturon on periphyton communities – a microcosm study. Estuarine, Coastal and Shelf Science, 62(3), 539545.
  • Schmitt-Jansen, M. & Altenburger, R. (2008) Community-level microalgal toxicity assessment by multiwavelength PAM fluorometry. Aquatic Toxicology, 86, 4958.
  • Schreiber, U., Müller, J.F., Haugg, A. & Gademann, R. (2002) New type of dual-channel PAM chlorophyll fluorometer for highly sensitive water toxicity biotests. Photosynthesis Research, 74, 317330.
  • Schreiber, U., Schliwa, U. & Bilger, W. (1986) Continuous recording of photochemical and non-photochemical chlorophyll fluorescence quenching with a new type of modulation fluorometer. Photosynthesis Research, 10, 5162.
  • Sijm, D., De Bruijn, J., Crommentuijn, T. & Van Leeuwen, K. (2001) Environmental quality standards: endpoints or triggers for a tiered ecological effect assessment approach? Environmental Toxicology and Chemistry, 20, 26442648.
  • Van der Grinten, E. (2005) Dynamic species interactions in phototrophic biofilms. PhD Thesis, University of Amsterdam, Amsterdam, The Netherlands.
  • Van der Heever, J.A. & Grobbelaar, J.U. (1996) Evaluation of a short-incubation-time small-volume radiocarbon-uptake algal toxicity test. Journal of Applied Phycology, 8, 6571.
  • Versteeg, D.J., Belanger, S.E. & Carr, G.J. (1999) Understanding single-species and model ecosystem sensitivity: data-based comparison. Environmental Toxicology and Chemistry, 18, 13291346.