Photoinduced Toxicity to Lake Erie Phytoplankton Assemblages from Intact and Photomodified Polycyclic Aromatic Hydrocarbons

J. Great Lakes Res. 29(4):558–565 Internat. Assoc. Great Lakes Res., 2003

Photoinduced Toxicity to Lake Erie Phytoplankton Assemblages from Intact and Photomodified Polycyclic Aromatic Hydrocarbons Christopher A. Marwood1, Ralph E.H. Smith1, Murray N. Charlton2, Keith R. Solomon3, and Bruce M. Greenberg1,* 1Department

of Biology, University of Waterloo Waterloo, Ontario N2L 3G1

2National

Water Research Institute P.O. Box 5050, Burlington Ontario L7R 4A6

3Centre

for Toxicology, University of Guelph Guelph, Ontario N1G 2W1

ABSTRACT. Pulse amplitude modulated (PAM) chlorophyll-a fluorescence is a simple, rapid technique for measuring photosynthetic efficiency in plants and algae that could be a useful biomarker of toxicity in the aquatic environment. PAM Chlorophyll fluorescence was used to detect inhibition of photosynthesis in natural assemblages of Lake Erie phytoplankton incubated with both intact and photomodified polycyclic aromatic hydrocarbons. The maximum efficiency of electron transport in photosystem II (Fv/Fm) and the effective yield of photochemistry (or steady state photosystem II activity) (µF/Fm′) were measured from phytoplankton exposed for 30 min in sunlight to polycyclic aromatic hydrocarbons at concentrations from 40 to 2,000 µgL–1. Anthracene, fluoranthene, and phenanthrenequinone were most toxic to phytoplankton, with EC50s for ∆F/Fm′ inhibition of 314, 118 and 90 µgL–1, respectively. Anthraquinone, 1,2dihydroxyanthraquinone, and phenanthrene were less toxic, with EC50s ranging from 684 to > 2,000 µgL–1. Recovery of photosynthetic function in darkness occurred to varying degrees, and was related to the regions of the photosynthetic apparatus on which the chemicals are thought to act. Inhibition of chlorophyll fluorescence parameters demonstrated a reciprocity-like response between concentration and duration of exposure, implying chronic exposures to lower concentrations of PAHs, such as those found in the Great Lakes, could cause inhibition of photosynthesis in phytoplankton assemblages. INDEX WORDS: Polycyclic aromatic hydrocarbons; ultraviolet radiation, phytoplankton, chlorophyll fluorescence, photosynthesis.

INTRODUCTION New tools are required to understand the impacts of environmental stresses and chemical contaminants on organisms in aquatic ecosystems. Although sensitive biological assays exist that can detect potential impacts, their use in the field has been limited. Rapid biomarkers of effects, which have been validated in the field, are required. Chlorophyll-a (Chl a) fluorescence induction is a simple technique for measuring photosynthetic efficiency in plants and algae. Measurements of photosynthetic efficiency are comparable to estimates of the rate of photosynthesis acquired with techniques *Corresponding

such as net CO2 fixation, but Chla fluorescence can be measured in a few minutes, rather than hours. Using a pulse-amplitude modulated (PAM) fluorometer, several parameters describing photosystem II (PSII) efficiency can be measured from a plant or an assemblage of phytoplankton cells with very little sample preparation. The maximum efficiency of electron transport through PSII is measured in the dark, and is described by the ratio Fv/Fm; where Fm is maximal fluorescence and Fv is Fm minus minimal fluorescence (Fo). The corresponding PSII efficiency in light-adapted plants or cells, ∆F/F m ′, represents the quantum yield of PSII during steadystate photosynthesis, and is closely correlated with rates of CO2 assimilation (Genty et al. 1989) and

author. E-mail: [email protected]

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PAH Toxicity to Phytoplankton oxygen evolution (Geel et al. 1997). ∆F/Fm′ corresponds to Fv/Fm but is measured during steady state photosynthesis (Bolhàr-Nordenkampf et al. 1989). The data acquired using Chla fluorescence have been found to be remarkably consistent across many different algae and plant groups, reflecting the conservation of photosynthetic processes throughout the plant kingdom (Bolhàr-Nordenkampf et al. 1989). Chla fluorescence is, however, sensitive to environmental stresses such as strong light, drought, low temperature, pathogen infection, and herbicide action (Lichtenthaler 1988, Huner et al. 1998). Because PSII is the first step in photosynthesis, Chla fluorescence potentially could be used to detect impacts of a number of different environmental toxicants and stressors that inhibit electron transport, carbon fixation, pigment production, or any other process associated with photosynthesis. An additional notable property of this assay is the ability to use it diagnostically, to discriminate effects of contaminants or stressors with different mechanisms of action (Brack and Frank 1998, Babu et al. 2001). Polycyclic aromatic hydrocarbons (PAHs) are common contaminants of aqueous environments (Cook et al. 1983, Eadie 1984). Sediment concentrations of PAHs in Lake Erie are higher in the west basin, due to influx from the highly contaminated Detroit River and adjacent industrialized areas (Eadie 1984, Metcalfe et al. 2000). Although most PAHs have low water solubility and remain bound to sediment, transport of PAHs adsorbed to particles (including plankton) in the water column occurs (Baker and Eisenreich 1989). Lake Erie phytoplankton are likely exposed to these chemicals as both move through the water column, albeit the PAH concentrations are usually low. Many PAHs inhibit photosynthesis and plant growth at low concentrations (Huang et al. 1997a, 1997b; Babu et al. 2001). Toxicity of many PAHs increases substantially in sunlight, due to production of reactive oxygen species upon absorbance of UVA light (Newsted and Giesy 1987). Absorbance of light also results in the formation of oxidized PAHs (oxyPAHs), which have been shown to have greater aqueous solubility and toxicity (Huang et al. 1997c, McConkey et al. 1997, Babu et al. 2001). However, the elevated toxicity of oxyPAHs is likely due to their mechanism of action rather than increased bioavailability (Duxbury et al. 1997, Huang et al. 1997b, Babu et al. 2001). We have previously shown that Chla fluorescence measured from natural assemblages of Lake Erie phytoplankton in vivo can be used to detect inhibition of photosynthesis by strong light (photoinhibi-

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tion) and by anthracene (Marwood et al. 1999, Marwood et al. 2000). In this study, Chla fluorescence was examined in phytoplankton exposed to three intact PAHs (anthracene, fluoranthene, and phenanthrene) commonly found in contaminated aquatic environments, as well as three oxyPAHs (anthraquinone, 1,2-dihydroxyanthraquinone, and 9,10phenanthrenequinone) that are formed in sunlight (Huang et al. 1997b, McConkey et al. 1997, Lampi et al. 2001). Assemblages of phytoplankton from several sites on Lake Erie were exposed in sunlight to each compound individually at various concentrations, allowing the generation of concentration-response curves and median toxicity estimators (EC50s). To extrapolate our results to lower, more environmentally relevant aqueous PAH concentrations, we tested for a reciprocity-like effect by varying inversely the exposure duration and concentration. To further examine the mechanism of toxicity, long term recovery of photosynthesis in darkness following sunlight exposure was also examined. METHODS AND MATERIALS Experiments were performed aboard C.C.G.S. Limnos (Canada Centre for Inland Waters, Burlington, Ontario) during survey cruises of Lake Erie in July, August, and September 1997. Due to the nature of the cruises, it was not possible to obtain replicate water samples. Instead, experiments using consistent techniques were repeated on different dates with samples from stations 974, 943, 946, 969, and 84 (Fig. 1). Whole water samples were taken early in the morning from a depth of 5 m using Niskin bottles and stored immediately in carboys in the dark.

FIG. 1. Lake Erie stations sampled for PAH exposure experiments. At each station, phytoplankton from 5m depth was sampled. Stations were sampled on different dates from June through August 1997.

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Stock solutions of 2 g L–1 of the intact PAHs anthracene (ANT), phenanthrene (PHE), and fluoranthene (FLA), and the oxyPAHs anthraquinone (ATQ), 9,10-phenanthrenequinone (PHEQ) and 1,2dihydroxyanthraquinone (1,2-dhATQ) (Sigma Chemical, St. Louis, MO, USA) were made by dissolving the pure chemicals in dimethylsulfoxide (Aldrich Chemical, Milwaukee, WI, USA). Chemicals were delivered to phytoplankton by diluting stock solutions to the appropriate concentration in 200 mL of lake water in polyethylene bags (First Brands, Scarborough, Ontario). The concentration of dimethylsulfoxide in dark and sunlight-exposed water samples was < 0.1% (v/v). Control samples without PAHs were incubated with 0.1% dimethylsulfoxide (v/v). We have previously shown that up to 1% (v/v) DMSO has no impact in the light or the dark on algal or plant photosynthesis (Huang et al. 1997a, 1997b; Marwood et al. 1999). Samples were incubated with individual chemicals for 30 min in the dark to allow uptake of the PAHs by phytoplankton, and then either kept in the dark or exposed to sunlight for 60 min on the deck of the ship in exposure chambers with circulating surface lake water to maintain a constant temperature. It has been shown previously that 30 min dark pre-exposures for up-take of PAHs only has minor impacts on inhibition of photosynthesis (Marwood et al. 1999). Exposure chambers were screened with polyester film (Mylar-D, 0.08 mm, Johnston Plastics) that removed ultraviolet-B (UVB, 290–320 nm) plus one layer of black mesh window screen that removed 50% of all wavelengths. This provided light levels similar to those at 5m depth in the central basin of Lake Erie; approximately 800µmol m–2s–1 of photosynthetically active radiation (PAR, 400 –700 nm) and 80 µmol m –2 s –1 of UVA (320–400 nm) (Smith et al. 1999). For recovery experiments, samples were transferred to dark chambers for 6h following sunlight exposure. Phytoplankton were concentrated onto 25mm glass fiber filters (GF/C; Whatman, Springfield Mill, UK) under low vacuum. Chl a fluorescence was measured directly from phytoplankton on moist filters floating on filtered lake filter. This technique has been found to yield consistent, reliable Chla fluorescence measurements from Lake Erie phytoplankton assemblages (Marwood et al. 1999). Chla fluorescence was measured using a pulseamplitude modulated fluorometer (PAM-2000, Walz, Effeltrich, Germany) from phytoplankton which had been dark-adapted for 15min. Initial flu-

orescence (Fo) was measured with a weak modulated red light (<1µmol m–2s–1). Dark-adapted maximum fluorescence (Fm) was induced by a single 0.6s saturating pulse of PAR (3,400 µmol m–2s–1). The short pulse of saturating light is well known not to cause any photoinhibition (Bolhàr-Nordenkampf et al. 1989, Marwood et al. 2000). Steady-state fluorescence (F s) and light-adapted maximum fluorescence (Fm′) were measured from phytoplankton after 10min in red actinic light (655nm, 22 µmol m–2s–1). The maximum efficiency of electron transport in photosystem II (Fv/Fm) and the effective yield of photosystem II photochemistry (∆F/F m ′) were calculated according to standard equations (Genty et al. 1989, van Kooten and Snel 1990) (Fig. 2): Fv/Fm = (Fm – Fo) / Fm

(1)

∆F/Fm′ = (Fm′ – Fs) / Fm′

(2)

Fv/Fm is maximal efficiency of photosystem II and represents the fraction of photosystem II units that are functional in a dark adapted algal sample. ∆F/F m′ represents the fraction of photosystem II units that are functional during steady-state electron transport (Bolhàr-Nordenkampf et al. 1989, Marwood et al. 2001b). Inhibition of Fv/Fm and ∆F/Fm′ was analyzed as a function of the nominal PAH concentration. Calculation of EC50s was performed by fitting the experimentally derived data to a logit function suitable for continuous response data (Sanathanan et al. 1987). Non-linear regression was used to fit Fv/Fm and ∆F/Fm′, expressed as the mean from three replicate experiments, to the equation % Inhibition = 100 / (1 + e β (x – µ))

(3)

where x was the logarithm of the concentration, µ was the logarithm of the EC50, and β represented the slope of the concentration-response curve. This gave a good estimate of the EC50 to allow a first approximation of the order of toxicity of the chemicals. RESULTS For most of the PAHs examined, short exposures of phytoplankton to PAHs in sunlight diminished Chla fluorescence. Exposure to PAHs diminished variable fluorescence (Fv = Fm–Fo), the difference between minimum and maximum fluorescence levels in the dark-adapted phytoplankton (Fig. 2). At

PAH Toxicity to Phytoplankton

FIG. 2. Effects of PAHs on Chla fluorescence in phytoplankton. Phytoplankton from station 974 were exposed to (A) 0, (B) 40, (C) 166, or (D) 500µgL–1 anthracene in sunlight for 30min. The minimum (Fo) and maximum (Fm) fluorescence measured in the dark are shown as the lower and upper solid lines; the minimum (Fs) and maximum (Fm′) fluorescence in the light are shown as the lower and upper dashed lines. the lowest PAH concentration tested (40µgL –1 ), maximum fluorescence in the light (Fm′) was diminished relative to maximum fluorescence in the dark (Fm). At higher concentrations, variable fluorescence was diminished nearly to zero, resulting in zero Fv/Fm and ∆F/Fm′ parameters as well.

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FIG. 3. Concentration-dependent inhibition of the yield of PSII photochemistry (∆F/Fm′) in phytoplankton exposed to intact PAHs and oxyPAHs. The plots represent samples from five different stations on Lake Erie, sampled on different dates. For each chemical, a logit function was fit to the data using non-linear regression to estimate the median toxic concentration (EC50). To keep the graphs from being to cluttered, only one representative experiment is shown from one site on Lake Erie.

Chla fluorescence was measured from several stations on Lake Erie on different dates. There was moderate variability in the data sets used to calculate EC50s, reflecting the different assemblages of phytoplankton at the various stations sampled. Nevertheless, for five of the six PAHs examined, concentration-response curves could be fit to the data by nonlinear regression. Both Fv/Fm and ∆F/Fm′ were diminished in a similar manner by PAHs in sunlight. For simplicity, only ∆F/Fm′ plots are shown (Fig. 3). The most toxic PAHs were the oxyPAH PHEQ, and the intact PAHs FLA and ANT, with ∆F/Fm′ EC50s ranging from 90 to 310 µgL–1 (Table 1). The other in-

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TABLE 1. Statistical parameters from regression curves for the inhibition of maximum efficiency of PSII electron transport (Fv /Fm) and yield of photochemistry (∆F/Fm′) in phytoplankton exposed to PAHs. Regression curves fit to the experimental data were described by a logit function, where µ was the logarithm of the PAH concentration (in µgL–1) at 50% inhibition and β was the slope of the regression curve. EC50 concentrations (± 95% confidence interval) were calculated from µ and β. Chemical ANT PHE FLA ATQ PHEQ 1,2-dhATQ

R2 0.978

µ 2.55

Fv/Fm β –1.363

0.785 0.774 0.962 0.911

2.10 2.82 2.02 3.09

–1.568 –1.147 –1.746 –2.321

EC50(µgL–1) 357 (±3.8) > 2,000 125 (±4.3) 654 (±8.8) 104 (±5.9) 1,229 (±4.9)

tact chemical, PHE, inhibited Chla fluorescence only moderately at all concentrations tested, and an EC50 could not be estimated. The remaining two oxyPAHs, ATQ and 1,2-dhATQ, inhibited Fv/Fm and ∆F/Fm′ only at the higher concentrations tested. For these chemicals, EC50s ranged from 654 to 1229 µgL–1. In general, EC50s for ∆F/Fm′ were slightly lower than Fv/Fm, except for ATQ. The PAH concentrations to which Lake Erie phytoplankton assemblages were exposed were higher than aqueous concentrations found in the Great Lakes, however the exposure duration of 30min was less than the amount of time phytoplankton are likely to spend in the photic zone of Lake Erie. A reciprocity-like effect between exposure concentration and duration was tested by measuring Chla fluorescence TABLE 2. Reciprocity–like effect between exposure duration and anthracene concentration. Lake Erie phytoplankton assemblages were exposed in sunlight to inversely proportional combinations of exposure duration and anthracene concentration. The maximum efficiency of PSII electron transport (Fv /Fm) and yield of photochemistry (∆F/Fm′) were measured using Chla fluorescence from phytoplankton isolated following exposure. Means from three replicate samples (± standard error) are shown. ANT Conc. (µgL–1) 0 37.5 75 150 300 600

Time (h) Control 4 2 1 0.5 0.25

Fv/Fm 0.372 (± 0.009) 0.222 (± 0.028) 0.145 (± 0.017) 0.151 (± 0.011) 0.123 (± 0.031) 0.135 (± 0.016)

∆F/Fm′ 0.265 (± 0.006) 0.141 (± 0.012) 0.091 (± 0.017) 0.086 (± 0.005) 0.065 (± 0.027) 0.068 (± 0.007)

R2 0.980

µ 2.50

∆F/Fm′ β –1.872

0.923 0.991 0.959 0.798

2.07 2.83 1.95 3.04

–2.250 –2.502 –2.849 –2.340

EC50(µgL–1) 314 (±2.7) > 2,000 118 (±2.6) 684 (±12.8) 90 (±2.4) 1,096 (±8.2)

from phytoplankton exposed to anthracene at different concentrations and duration, such that the product of the concentration and time was constant. Both Fv/Fm and ∆F/Fm′ were diminished to a similar degree by short exposures to high anthracene concentrations, and by long exposures to low concentrations (Table 2). Only at the longest duration and lowest concentration were Fv/Fm and ∆F/Fm′ significantly greater than the other combinations tested. The ability of phytoplankton to recover photosynthetic function following exposure to PAHs in sunlight was examined. F v /F m and ∆F/F m ′ were measured from samples 6 h after a 30min exposure in sunlight to 500 µgL–1 of a given chemical. During the recovery period, phytoplankton were kept in darkness to minimize photoinduced toxicity. For all PAHs except PHEQ, phytoplankton demonstrated some recovery of Fv/Fm in darkness (Table 3). The degree of recovery depended on the chemical to which they were exposed and was consistent with the putative mechanisms of action of the different chemicals (see discussion). Exposure to PHE diminished Fv/Fm and ∆F/Fm′ to 60% of control levels, but both parameters recovered to 90% of control values. Phytoplankton exposed to ANT, FLA, ATQ, and 1,2-dhATQ recovered F v /F m by only 20–30% during the 6-hour dark period. Interestingly, ∆F/Fm′ in samples exposed to ANT, ATQ and 1,2-dhATQ did not recover, but showed further diminishment during the period in darkness. In plankton exposed to PHEQ, both Fv/Fm and ∆F/Fm′ were severely diminished to only 5% of control values, and did not demonstrate recovery. DISCUSSION Chla fluorescence is a potentially powerful biomarker of environmental contaminants in aquatic

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TABLE 3. Recovery of photosynthesis in phytoplankton assemblages in darkness following exposure to PAHs in sunlight. The maximum efficiency of PSII electron transport (Fv/Fm) and yield of photochemistry (∆F/Fm′) were measured using Chla fluorescence from phytoplankton following a 30 min exposure in sunlight to 500µgL–1 of chemical. The same parameters were measured from samples incubated in sunlight followed by 6h in darkness to allow recovery from photoinduced toxicity. Chemical Control ANT PHE FLA ATQ PHEQ 1,2-dhATQ

Fv/Fm 0.291 0.103 0.179 0.066 0.108 0.015 0.160

0.5 h exposure % Control ∆F/Fm′ % Control — 0.191 — 35 0.051 27 62 0.124 65 23 0.028 15 37 0.086 45 5 0.011 6 55 0.155 81

ecosystems. The information provided by this assay is comparable to traditional techniques for assessing photosynthesis, such as CO2-uptake and oxygen evolution. The parameter µF/F m ′ estimates PSII photosynthetic yield of electron transport (φPSII), and is strongly correlated with steady state CO2 assimilation (Genty et al. 1989). This parameter integrates effects on many processes that contribute to inhibition of electron transport, either by decreasing the quantum yield of photochemistry or by increasing the yield of non-photochemical, thermal decay processes. In contrast with traditional techniques, Chla fluorescence can detect inhibition of photosynthesis in minutes rather than hours, with relatively little sample preparation. These assay properties are essential in studies of rapid or transient effects, such as the experiments described here. Chla fluorescence has been used previously to detect diminished photosynthesis in Lake Erie phytoplankton assemblages from solar UVB radiation (Marwood et al. 2000). To eliminate the direct effects of sunlight on photosynthesis, UVB wavelengths were removed with filters. As well, potential photoinhibition from high levels of PAR and UVA wavelengths was eliminated by attenuating sunlight by half during the exposures. The light treatment used in the PAH experiments was sufficient to generate photoinduced toxicity from PAHs, but had minor effects on photosynthesis in the absence of PAHs. The light levels received by phytoplankton in the exposure chambers (~800 µmol m –2s –1 of PAR and 80 µmol m –2s –1 of UVA) approximated the levels at 2 to 5m depth in Lake Erie, depending on location (Smith et al. 1999). Of the intact PAHs examined in this study, the greatest toxicity in sunlight was from ANT and

Fv/Fm — 0.192 0.261 0.117 0.130 0.006 0.256

0.5 h exposure + 6 h darkness % Control ∆F/Fm′ % Control — — — 66 0.043 23 90 0.170 89 40 0.096 50 45 0.057 30 2 0.009 5 88 0.149 78

FLA, which are both efficient photosensitizers capable of generating reactive oxygen species that oxidize biomolecules (Huang et al. 1997c) and indirectly disrupt cellular processes such as electron transport. The least toxic PAH was PHE, which is predicted by QSAR models to be a poor photosensitizer (Newsted and Giesy 1987, Huang et al. 1997c). Diminished maximum photochemical efficiency of PSII (Fv/Fm) could indicate degradation of the proteins in the PSII reaction centers, resulting in a large proportion of inactive PSII. The oxyPAHs examined in this study are poor photosensitizers, but many have been shown to act as electron acceptors and directly inhibit photosynthetic electron transport (Oettmeier et al. 1988, Huang et al. 1997b, Babu et al. 2001). The greatest toxicity to Lake Erie phytoplankton was from PHEQ, which is toxic at similar concentrations to the marine bacteria Vibrio fischeri (EC50 = 60 µgL–1) and the duckweed Lemna gibba (EC50 = 43 µgL –1 ) (McConkey et al. 1997). Toxicity from PHEQ is independent of light, suggesting that, unlike the other PAHs examined in this study, PHEQ is a poor photosensitizer. Alternatively, PHEQ may act as a redox cycler, disrupting the photosynthetic electron transport chain by removing electrons from the electron transport chain downstream of PSII. Some quinones with chemical structures similar to the intermediate carrier plastoquinone are known to disrupt electron transport (Oettmeier et al. 1988). The lack of recovery in darkness of phytoplankton exposed to PHEQ supports such a mechanism of action. Chl a fluorescence is capable of detecting this type of disruption as well as direct effects on PSII. Because PSII is the first step in photosynthetic electron transport, inhibition of electron

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transport at any point within the electron transport chain exerts excitation pressure on PSII, and can be detected as diminished Fv/Fm and ∆F/Fm′ (Huang et al. 1997b, Huner et al. 1998). The concentrations of PAHs which were toxic to phytoplankton in our short exposures were within the range of concentrations found in water overlying highly polluted sediments (Fallon and Horvath 1985, Murphy et al. 1990), but were higher than aqueous levels found in most parts of Lake Erie (Eadie 1984). However, the reciprocity-like effect between concentration and exposure duration observed in this study indicates that short exposures to high concentrations are suitable for predicting impacts on phytoplankton by lower PAH concentrations and longer exposure conditions, such as those experienced by natural assemblages of Lake Erie phytoplankton. For example, a concentration of 75 µgL–1 ANT is still within the range of aqueous PAH concentrations found in contaminated areas of the Great Lakes (Eadie 1984, Furlong et al. 1988). Our experiments indicate that phytoplankton exposed to this PAH concentration for only 2h had substantially diminished photosynthesis. The inability to recover photosynthetic capacity following exposure to many of the PAHs tested in this study, even after prolonged periods in darkness, suggests inhibition of photosynthesis can be extensive following even transient exposure to PAHs. This indicated that phytoplankton even in mildly contaminated sites of Lake Erie may be affected by PAHs and oxyPAHs present in the water column, even at moderate or low chemical concentrations. The ability of Chla fluorescence to detect inhibition of photosynthesis from a variety of PAHs with distinctly different mechanisms of toxicity indicates this assay can be a general biomarker of stressor effects. However, we have found that effects of different environmental stressors, such as UVB radiation, can be distinguished from PAH toxicity in phytoplankton, and in freshwater plants, using Chla fluorescence. In plants exposed to PAHs, variable fluorescence (Fv) is often diminished, suggesting possible direct effects on electron transport, due to blocked PSII or diversion of electrons (Marwood et al. 1999, Marwood et al. 2001a). Strong sunlight containing UVB wavelengths, on the other hand, lowers the fluorescence signal F o as well as F v, which is attributed to loss of Chl and transfer of energy to xanthophyll pigments (Marwood et al. 2000, Marwood et al. 2001b). In addition to PAHs and UV radiation, effects of contaminants such as triazine herbicides and pentachlorophenol may be

distinguished, based on the relative inhibition of fluorescence parameters associated with specific photosynthetic processes (Brack and Frank 1998). When applied appropriately, the Chla fluorescence technique may be used in a diagnostic manner to differentiate the toxic effects in plants and algae exposed to multiple environmental contaminants and stressors in the field. ACKNOWLEDGMENTS This work was supported by grants from the Canadian Network of Toxicology Centres (B.M.G., K.R.S.), the Natural Sciences and Engineering Research Council (B.M.G.), and Science Horizons / Environment Canada. REFERENCES Babu, T.S., Marder, J.B., Tripuranthankan, S., Dixon, D. G., and Greenberg, B.M. 2001. Synergistic effects of a photooxidized PAH and copper on photosynthesis and plant growth: Evidence that active oxygen formation is a mechanism of copper toxicity. Environ. Toxicol. Chem. 20:1351–1358. Baker, J.E., and Eisenreich, S.J. 1989. PCBs and PAHs as tracers of particulate dynamics in the large lakes. J. Great Lakes Res. 15:84–103. Bolhàr-Nordenkampf, H.R., Long, S.P., Baker, N.R., Öquist, G., Schreiber, U., and Lechner, E.G. 1989. Chlorophyll fluorescence as a probe of the photosynthetic competence of leaves in the field: a review of current instrumentation. Funct. Ecol. 3:497–514. Brack, W., and Frank, H. 1998. Chlorophyll a fluorescence: a tool for the investigation of toxic effects in the photosynthetic apparatus. Ecotoxicol. Environ. Saf. 40:34–41. Cook, D.K., Pierce, R.C., Eaton, P.B., Lao, R.C., Onuska, F.I., Payne, J.F., and Vavasour, E. 1983. Polycyclic aromatic hydrocarbons in the aquatic environment: formation, sources, fate and effects on aquatic biota. National Research Council of Canada, Ottawa, Ontario, Canada. NRCC 18981. Duxbury, C.L., Dixon, D.G., and Greenberg, B.M. 1997 The effects of simulated solar radiation on the bioaccumulation of polycyclic aromatic hydrocarbons by the duckweed Lemna gibba. Environ. Toxicol. Chem. 16:1739–1748. Eadie, B.J. 1984. Distribution of polycyclic aromatic hydrocarbons in the Great Lakes. In Advances in Environmental Science and Technology, eds. J.O. Nriagu and M.S. Simmons, pp. 195–211. New York, NY, USA: John Wiley and Sons. Fallon, M.E., and Horvath, F.J. 1985. Preliminary assessment of contaminants in soft sediments of the Detroit River. J. Great Lakes Res. 11:373–378. Furlong, E.T., Carter, D.S., and Hites, R.A. 1988.

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