# Resistance and recovery of river biofilms receiving short pulses of Triclosan and Diuron

Lorenzo Proia, Soizic Morin, M. Peipoch, A.M. Romani, Sergi Sabater

► To cite this version:

Lorenzo Proia, Soizic Morin, M. Peipoch, A.M. Romani, Sergi Sabater. Resistance and recovery of river biofilms receiving short pulses of Triclosan and Diuron. Science of the Total Environment, 2011, 409 (17), pp.3129-3137. 10.1016/j.scitotenv.2011.05.013. hal-02595139

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1 RESISTANCE AND RECOVERY OF RIVER BIOFILMS RECEIVING SHORT

2 PULSES OF TRICLOSAN AND DIURON

3

4 L. Proia¹, Morin S.¹/², Peipoch M.¹/³, Romaní A.M.¹, S. Sabater ¹/⁴

5

6 ¹Institute of Aquatic Ecology, University of Girona, E-17071 Girona (Spain)

7 ²Cemagref, UR REBX, F-33612 Cestas Cedex (France)

8 ³Centre d'Estudis Avançats de Blanes, Consejo Superior de Investigaciones Científicas, E-17300 Blanes (Spain)

9 ⁴Catalan Institute for Water Research (ICRA), E-17003 Girona (Spain)
1 RESISTANCE AND RECOVERY OF RIVER BIOFILMS RECEIVING SHORT

2 PULSES OF TRICLOSAN AND DIURON

3

4 L. Proia¹, Morin S.¹/², Peipoch M.¹/³, Romaní A.M.¹, S. Sabater ¹/⁴

5

6 ¹Institute of Aquatic Ecology, University of Girona, E-17071 Girona (Spain)

7 ²Cemagref, UR REBX, F-33612 Cestas Cedex (France)

8 ³Centre d'Estudis Avançats de Blanes, Consejo Superior de Investigaciones Científicas, E-17300 Blanes (Spain)

9 ⁴Catalan Institute for Water Research (ICRA), E-17003 Girona (Spain)
13 ABSTRACT

14 The effects of the herbicide Diuron (DIU) and the bactericide Triclosan (TCS) were assessed on
15 laboratory-grown stream biofilms. Four week-old biofilms were exposed in mesocosms to 48-
16 hours of short pulses of either DIU or TCS. The direct and indirect effects of each toxicant on
17 the biofilms, and the subsequent recovery of the biofilms, were evaluated according to structural
18 and functional biomarkers. These parameters were analyzed immediately before exposure,
19 immediately after exposure, and 9 and 16 days post-exposure. DIU caused an increase in diatom
20 mortality (+79%), which persisted until the end of the experiment. TCS also affected diatom
21 mortality (+41%), although the effect did not appear until one week post-exposure. TCS caused
22 an increase in bacterial mortality (+45%); however, this parameter returned to normal values 1
23 week post-exposure. TCS compromised the cellular integrity of the green alga *Spirogyra* sp.,
24 whereas DIU did not. TCS also strongly inhibited phosphate uptake (-71%), which did not
25 return to normal values until 2 weeks post-exposure. DIU directly affected algae, but barely
26 affected the heterotrophs, whereas TCS seriously impaired bacteria (direct effect) as well as
27 autotrophs (indirect effect). However, the biofilms recovered their normal structure and function
28 within only a few days to a few weeks. These findings demonstrate the capacity of biofilms to
29 cope with periodic inputs of toxicants, but also the risks associated to repeated exposure or
30 multi-contamination in aquatic ecosystems.

31 Keywords:

32 Biofilms, Resistance, Recovery, Triclosan, Diuron, Pulses.

33

34
13 ABSTRACT

14 The effects of the herbicide Diuron (DIU) and the bactericide Triclosan (TCS) were assessed on
15 laboratory-grown stream biofilms. Four week-old biofilms were exposed in mesocosms to 48-
16 hours of short pulses of either DIU or TCS. The direct and indirect effects of each toxicant on
17 the biofilms, and the subsequent recovery of the biofilms, were evaluated according to structural
18 and functional biomarkers. These parameters were analyzed immediately before exposure,
19 immediately after exposure, and 9 and 16 days post-exposure. DIU caused an increase in diatom
20 mortality (+79%), which persisted until the end of the experiment. TCS also affected diatom
21 mortality (+41%), although the effect did not appear until one week post-exposure. TCS caused
22 an increase in bacterial mortality (+45%); however, this parameter returned to normal values 1
23 week post-exposure. TCS compromised the cellular integrity of the green alga *Spirogyra* sp.,
24 whereas DIU did not. TCS also strongly inhibited phosphate uptake (-71%), which did not
25 return to normal values until 2 weeks post-exposure. DIU directly affected algae, but barely
26 affected the heterotrophs, whereas TCS seriously impaired bacteria (direct effect) as well as
27 autotrophs (indirect effect). However, the biofilms recovered their normal structure and function
28 within only a few days to a few weeks. These findings demonstrate the capacity of biofilms to
29 cope with periodic inputs of toxicants, but also the risks associated to repeated exposure or
30 multi-contamination in aquatic ecosystems.

31 Keywords:

32 Biofilms, Resistance, Recovery, Triclosan, Diuron, Pulses.

33

34
35

# 1. INTRODUCTION

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37

38

Pollutants from agricultural, industrial and domestic activities enter watercourses either continuously (producing potentially chronic effects) or in pulses (causing potentially acute effects), in function of flow episodes, crop treatments and/or industrial release (Ellis, 2006).

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These chronic and periodic inputs are likely to have unexpected effects on the organisms living in aquatic environments. River ecosystems feature various ecological services (e.g. self-depuration and organic matter mineralization) directly related to processes driven by complex microbial communities (Mathuriau and Chauvet, 2002; Findlay et al., 1993; Sabater et al., 2007). These communities include benthic biofilms, which play a fundamental role in the trophic web and in the geochemical cycles within aquatic ecosystems (Battin et al., 2003; Lock, 1993). As interfaces between the water column and the substrata, biofilms are the first communities to suffer the consequences of pollutants (Sabater et al., 2007). Thus, understanding the resistance and resilience of biofilm communities to pollutants is crucial for ecological risk assessment of priority and emerging compounds.

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This study analyzes the effects of two compounds on stream biofilms: the herbicide Diuron (DIU; 3-(3,4-dichlorophenyl)-1,1-dimethylurea), and the broad-spectrum bactericide Triclosan (TCS; 5-chloro-2-(2,4-dichlorophenoxy)phenol), which operate by different modes of action.

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DIU is a photosynthesis inhibitor included on the list of priority pollutants of the EU Water Framework Directive (European Commission, 2000). As herbicide, DIU is active against phototrophic microorganisms and higher plants by blocking the chloroplast electron transport chain in Photosystem-II (Moreland, 1967). It has been used to control various annual and perennial broadleaf and grassy weeds, and is applied for vineyard protection. It has been also used on non-crop areas such as roads, garden paths and railway lines, and on many agricultural crops such as fruit, cotton, sugar cane, alfalfa and wheat (Giacomazzi and Cochet, 2004). Several
35

# 1. INTRODUCTION

36

37

38

Pollutants from agricultural, industrial and domestic activities enter watercourses either continuously (producing potentially chronic effects) or in pulses (causing potentially acute effects), in function of flow episodes, crop treatments and/or industrial release (Ellis, 2006).

39

40

41

42

43

44

45

46

47

48

These chronic and periodic inputs are likely to have unexpected effects on the organisms living in aquatic environments. River ecosystems feature various ecological services (e.g. self-depuration and organic matter mineralization) directly related to processes driven by complex microbial communities (Mathuriau and Chauvet, 2002; Findlay et al., 1993; Sabater et al., 2007). These communities include benthic biofilms, which play a fundamental role in the trophic web and in the geochemical cycles within aquatic ecosystems (Battin et al., 2003; Lock, 1993). As interfaces between the water column and the substrata, biofilms are the first communities to suffer the consequences of pollutants (Sabater et al., 2007). Thus, understanding the resistance and resilience of biofilm communities to pollutants is crucial for ecological risk assessment of priority and emerging compounds.

49

50

51

52

This study analyzes the effects of two compounds on stream biofilms: the herbicide Diuron (DIU; 3-(3,4-dichlorophenyl)-1,1-dimethylurea), and the broad-spectrum bactericide Triclosan (TCS; 5-chloro-2-(2,4-dichlorophenoxy)phenol), which operate by different modes of action.

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56

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59

DIU is a photosynthesis inhibitor included on the list of priority pollutants of the EU Water Framework Directive (European Commission, 2000). As herbicide, DIU is active against phototrophic microorganisms and higher plants by blocking the chloroplast electron transport chain in Photosystem-II (Moreland, 1967). It has been used to control various annual and perennial broadleaf and grassy weeds, and is applied for vineyard protection. It has been also used on non-crop areas such as roads, garden paths and railway lines, and on many agricultural crops such as fruit, cotton, sugar cane, alfalfa and wheat (Giacomazzi and Cochet, 2004). Several
studies have reported the presence of DIU in surface waters (Azevedo et al., 2000; Blanchoud et al. 2004; Rodriguez-Mozaz et al., 2004).

TCS is active against both gram-positive and gram-negative bacteria. It is an inhibitor of the enzyme enoyl-acyl carrier protein reductase (ENR), which is involved in bacterial lipid biosynthesis (Adolfsson-Erici et al., 2002). For over 30 years TCS has been used in products such as anti-bacterial hand soaps, deodorants, household cleaners, dental hygiene products, and textiles (Singer et al., 2002). This emerging compound has been reported in sewage wastewater and sludge at significant concentrations (Halden and Paull, 2005; Samsøe-Petersen et al., 2003). Although wastewater treatment plants (WWTPs) are rather effective at removing TCS (Samsøe-Petersen et al., 2003; McAvoy et al., 2002), this compound still reaches freshwater systems, and has been reported in various aquatic habitats, including rivers, streams (Ellis, 2006; Kuster et al., 2008; Morral et al., 2004), lakes (Loos et al., 2007; Singer et al., 2002) and the sea (Xie et al., 2008). Both DIU and TCS have been widely tested for toxicity to myriad cultured aquatic organisms (Canesi et al., 2007; Capdevielle et al., 2008; De Lorenzo et al., 2007; Farré et al., 2008; Flaherty and Dodson, 2005; Giacomazzi and Cochet, 2004; Orvos et al., 2002; Wilson et al., 2003); however, they have not been extensively studied for toxicity to natural complex communities (Franz et al., 2008; Lawrence et al., 2009; Morin et al., 2010a; Pesce et al., 2006, 2008; Ricart et al., 2009).

DIU reaches running waters primarily via pulses from diffuse source, whereas TCS enters them periodically from fixed sources (chiefly, WWTPs). DIU pulses of up to 134.0 µg L⁻¹ have been described during flooding events in vineyard catchments. These chronically affected environments show baseline concentrations of 1 µg DIU L⁻¹ between flood events (Rabiet et al., 2010). In contrast, TCS enters running waters chronically at low concentrations via urban sewage effluents, and its removal during wastewater treatment is variable (Ellis, 2006; Ricart et al., 2010). Measured TCS concentrations can be reduced to 80% on average through waste water
studies have reported the presence of DIU in surface waters (Azevedo et al., 2000; Blanchoud et al. 2004; Rodriguez-Mozaz et al., 2004).

TCS is active against both gram-positive and gram-negative bacteria. It is an inhibitor of the enzyme enoyl-acyl carrier protein reductase (ENR), which is involved in bacterial lipid biosynthesis (Adolfsson-Erici et al., 2002). For over 30 years TCS has been used in products such as anti-bacterial hand soaps, deodorants, household cleaners, dental hygiene products, and textiles (Singer et al., 2002). This emerging compound has been reported in sewage wastewater and sludge at significant concentrations (Halden and Paull, 2005; Samsøe-Petersen et al., 2003). Although wastewater treatment plants (WWTPs) are rather effective at removing TCS (Samsøe-Petersen et al., 2003; McAvoy et al., 2002), this compound still reaches freshwater systems, and has been reported in various aquatic habitats, including rivers, streams (Ellis, 2006; Kuster et al., 2008; Morral et al., 2004), lakes (Loos et al., 2007; Singer et al., 2002) and the sea (Xie et al., 2008). Both DIU and TCS have been widely tested for toxicity to myriad cultured aquatic organisms (Canesi et al., 2007; Capdevielle et al., 2008; De Lorenzo et al., 2007; Farré et al., 2008; Flaherty and Dodson, 2005; Giacomazzi and Cochet, 2004; Orvos et al., 2002; Wilson et al., 2003); however, they have not been extensively studied for toxicity to natural complex communities (Franz et al., 2008; Lawrence et al., 2009; Morin et al., 2010a; Pesce et al., 2006, 2008; Ricart et al., 2009).

DIU reaches running waters primarily via pulses from diffuse source, whereas TCS enters them periodically from fixed sources (chiefly, WWTPs). DIU pulses of up to 134.0 µg L⁻¹ have been described during flooding events in vineyard catchments. These chronically affected environments show baseline concentrations of 1 µg DIU L⁻¹ between flood events (Rabiet et al., 2010). In contrast, TCS enters running waters chronically at low concentrations via urban sewage effluents, and its removal during wastewater treatment is variable (Ellis, 2006; Ricart et al., 2010). Measured TCS concentrations can be reduced to 80% on average through waste water
treatment plants (Kantiani et al., 2008), but still can reach river waters, where observed
concentration range between 0.027 and 2.7µg L⁻¹ (Ricart et al., 2009). In spite of these low
concentrations, brief spikes of compounds such as TCS could occur during dry periods, and
might result in transient perturbations of river ecosystems, with unknown long-term implications.
These inputs may have specific or non-specific effects on both target and non-target organisms.
Studying how biological communities are affected from these events is an ecological priority.
Studying how communities recover from transient perturbations is important for assessing the
risks associated with chronic contaminations

This study gauged the ability of biofilms to cope with short pulses of either DIU or TCS, assessing their initial responses and their subsequent recoveries. Considering the intrinsic complexity of biofilms, it was hypothesized that in addition to the specific effects of DIU on autotrophs, and of TCS on bacteria, these toxicants could provoke indirect effects deriving from ecological interactions at the microbial scale. The direct effects were expected to occur immediately upon toxicant exposure, whereas the indirect ones were expected to appear later on. Moreover, it was predicted that the time required for the biofilms to recover from the pollutant pulse could be correlated to the presence of target organisms (direct effects) and non-target organisms (indirect effects) in the biofilms. Given that indirect effects can involve much more complex mechanisms (Ricart et al., 2009), they can imply longer recovery times. Thus, the core hypothesis in this work was that early recovery of biofilm endpoints could be related to direct effects of pollutants, whereas late recovery could be related to indirect ones. In order to verify our hypothesis toxicants concentrations were selected after considering ecotoxicological data available (i.e. EC50, NOEC) both on single cultured species and on natural biofilm communities. It has been demonstrated how complex biofilm communities can result more resistant to toxicants than single species composing it (Franz et al., 2008). Several studies demonstrated the protective function of extracellular polymeric substances produced by biofilm organisms (Admiraal et al.,
treatment plants (Kantiani et al., 2008), but still can reach river waters, where observed concentration range between 0.027 and 2.7µg L⁻¹ (Ricart et al., 2009). In spite of these low concentrations, brief spikes of compounds such as TCS could occur during dry periods, and might result in transient perturbations of river ecosystems, with unknown long-term implications. These inputs may have specific or non-specific effects on both target and non-target organisms. Studying how biological communities are affected from these events is an ecological priority. Studying how communities recover from transient perturbations is important for assessing the risks associated with chronic contaminations

This study gauged the ability of biofilms to cope with short pulses of either DIU or TCS, assessing their initial responses and their subsequent recoveries. Considering the intrinsic complexity of biofilms, it was hypothesized that in addition to the specific effects of DIU on autotrophs, and of TCS on bacteria, these toxicants could provoke indirect effects deriving from ecological interactions at the microbial scale. The direct effects were expected to occur immediately upon toxicant exposure, whereas the indirect ones were expected to appear later on. Moreover, it was predicted that the time required for the biofilms to recover from the pollutant pulse could be correlated to the presence of target organisms (direct effects) and non-target organisms (indirect effects) in the biofilms. Given that indirect effects can involve much more complex mechanisms (Ricart *et al.*, 2009), they can imply longer recovery times. Thus, the core hypothesis in this work was that early recovery of biofilm endpoints could be related to direct effects of pollutants, whereas late recovery could be related to indirect ones. In order to verify our hypothesis toxicants concentrations were selected after considering ecotoxicological data available (i.e. EC50, NOEC) both on single cultured species and on natural biofilm communities. It has been demonstrated how complex biofilm communities can result more resistant to toxicants than single species composing it (Franz et al., 2008). Several studies demonstrated the protective function of extracellular polymeric substances produced by biofilm organisms (Admiraal *et al.*,
110 1999; Samrakandi *et al.*, 1997). These evidences were considered for the selection of toxicants concentrations as well as for consequent comparison of results with other studies.

112 In this work, 4-week old biofilms were subjected to 48-hours of short pulses of either DIU or
113 TCS. Their responses to these compounds were measured just after exposure, and then once
114 weekly for 2 weeks post-exposure. Each toxicant's effects on the biomass and survival of algae
115 and bacteria were used to gauge their respective effects on biofilm structure. Furthermore, their
116 effects on extracellular enzymatic activity and phosphate uptake were used to assess the
117 ecological implications of their entry into running water—namely, in the context of nutrient
118 retention and river self-depuration.

119

120 **2. MATERIAL AND METHODS**

2.1 Experimental design

122 Biofilms were scraped from rocks of the Fuirosos Stream, a third order pristine stream located in
123 the Natural Park of Montnegre-Corredor, (50 km N of Barcelona, NE Spain), and then inoculated
124 and colonized in twelve independent mesocosms. They were colonized on glass slides (1 cm²
125 each) placed at the bottom of each mesocosm (35-40 slides per mesocosm).The mesocosms
126 comprised sterile glass jars (19 cm in diameter, 9 cm high), filled with 1.5 L of artificial stream
127 water which was recirculated using a submersible pump (Hydor, Pico 300, 230 V 50 Hz, 4.5W).

128 Artificial stream water was produced adding pure salts to MilliQ water (Millipore) as described in
129 Ylla *et al.* (2009). To avoid nutrient depletion the water in each mesocosm was changed twice
130 weekly. All mesocosms were maintained in an incubator (Radiber AGP-570) under controlled
131 temperature (17.5 ± 1.1 °C) and light irradiance (160-180 µmol photons m⁻² s⁻¹; darkness/light
132 cycle of 12 h/12 h). After 4 weeks colonization, four mesocosms were treated with TCS
133 (IRGASAN, Sigma Aldrich, >97%, CAS: 3380-34-5) up to nominal concentrations of 60 µg/L,
134 and another four, with DIU (Sigma Aldrich, minimum 98%, CAS: 330-54-1) up to nominal
110 1999; Samrakandi *et al.*, 1997). These evidences were considered for the selection of toxicants
111 concentrations as well as for consequent comparison of results with other studies.

112 In this work, 4-week old biofilms were subjected to 48-hours of short pulses of either DIU or
113 TCS. Their responses to these compounds were measured just after exposure, and then once
114 weekly for 2 weeks post-exposure. Each toxicant's effects on the biomass and survival of algae
115 and bacteria were used to gauge their respective effects on biofilm structure. Furthermore, their
116 effects on extracellular enzymatic activity and phosphate uptake were used to assess the
117 ecological implications of their entry into running water—namely, in the context of nutrient
118 retention and river self-depuration.
119

120

## 2. MATERIAL AND METHODS

120

### 2.1 Experimental design

122 Biofilms were scraped from rocks of the Fuirosos Stream, a third order pristine stream located in
123 the Natural Park of Montnegre-Corredor, (50 km N of Barcelona, NE Spain), and then inoculated
124 and colonized in twelve independent mesocosms. They were colonized on glass slides (1 cm²
125 each) placed at the bottom of each mesocosm (35-40 slides per mesocosm). The mesocosms
126 comprised sterile glass jars (19 cm in diameter, 9 cm high), filled with 1.5 L of artificial stream
127 water which was recirculated using a submersible pump (Hydor, Pico 300, 230 V 50 Hz, 4.5W).
128 Artificial stream water was produced adding pure salts to MilliQ water (Millipore) as described in
129 Ylla *et al.* (2009). To avoid nutrient depletion the water in each mesocosm was changed twice
130 weekly. All mesocosms were maintained in an incubator (Radiber AGP-570) under controlled
131 temperature (17.5 ± 1.1 °C) and light irradiance (160-180 µmol photons m⁻² s⁻¹; darkness/light
132 cycle of 12 h/12 h). After 4 weeks colonization, four mesocosms were treated with TCS
133 (IRGASAN, Sigma Aldrich, >97%, CAS: 3380-34-5) up to nominal concentrations of 60 µg/L,
134 and another four, with DIU (Sigma Aldrich, minimum 98%, CAS: 330-54-1) up to nominal

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concentrations of 15 µg/L. The last four mesocosms were left untreated and used as control. To minimize photodegradation of the toxicants and ionization of TCS (pKa = 8.1), water and toxicants were renewed every 3 h (during the light cycle) for 48 hours. The pH was monitored between water changes during toxicant exposure. After the exposure period, the mesocosms were refilled with unpolluted artificial river water (as described above) that was changed twice per week during the following 2 weeks. The biofilm was sampled four times: before contamination (day 0), after the 48-hour exposure (day 2), and 1 and 2 weeks post-exposure (days 9 and 16).

Glass tiles from each mesocosm were randomly sampled. Extracellular enzyme activities, photosynthetic parameters ($F_0$, $Y_{eff}$ and $Y_{max}$), and bacterial densities were immediately measured after collection. Diatom samples for enumeration and taxonomical identification were preserved in formalin before being processed. Samples for chlorophyll determination were frozen (-20 °C) until analysis. Samples for Scanning Electron Microscopy (SEM) observation were collected at day 16. The phosphorus uptake (P-uptake) of the biofilm in each mesocosm was experimentally determined on each sampling day.

## 2.2 Water analysis

Concentrations of TCS, methyl-Triclosan (Me-TCS) and DIU were determined using high performance liquid chromatography (HPLC). Stock solutions (1 mg/mL) were prepared by dissolving pure standards of the highest purity available (HPLC grade, Sigma Aldrich) in methanol. An external calibration curve was then built for each compound by injecting different concentrations of individual standards prepared by different dilutions of the stock solution.

Water samples were collected once from each mesocosm. Samples were filtered through 0.45µm nylon membrane filters (Whatman) and immediately loaded onto C18 SPE cartridges (Sep-Pak® Vac 3 cc tC18, Waters, Ireland) previously conditioned with 5 mL of HPLC water and methanol at a flow rate of 1 mL min⁻¹. Samples (500 mL) were loaded at a flow rate of 5 mL min⁻¹. After pre-concentration, the cartridges were completely dried in vacuo for 20 minutes
concentrations of 15 µg/L. The last four mesocosms were left untreated and used as control. To minimize photodegradation of the toxicants and ionization of TCS (pKa = 8.1), water and toxicants were renewed every 3 h (during the light cycle) for 48 hours. The pH was monitored between water changes during toxicant exposure. After the exposure period, the mesocosms were refilled with unpolluted artificial river water (as described above) that was changed twice per week during the following 2 weeks. The biofilm was sampled four times: before contamination (day 0), after the 48-hour exposure (day 2), and 1 and 2 weeks post-exposure (days 9 and 16).

Glass tiles from each mesocosm were randomly sampled. Extracellular enzyme activities, photosynthetic parameters ($F_0$, $Y_{eff}$ and $Y_{max}$), and bacterial densities were immediately measured after collection. Diatom samples for enumeration and taxonomical identification were preserved in formalin before being processed. Samples for chlorophyll determination were frozen (-20 °C) until analysis. Samples for Scanning Electron Microscopy (SEM) observation were collected at day 16. The phosphorus uptake (P-uptake) of the biofilm in each mesocosm was experimentally determined on each sampling day.

## 2.2 Water analysis

Concentrations of TCS, methyl-Triclosan (Me-TCS) and DIU were determined using high performance liquid chromatography (HPLC). Stock solutions (1 mg/mL) were prepared by dissolving pure standards of the highest purity available (HPLC grade, Sigma Aldrich) in methanol. An external calibration curve was then built for each compound by injecting different concentrations of individual standards prepared by different dilutions of the stock solution.

Water samples were collected once from each mesocosm. Samples were filtered through 0.45µm nylon membrane filters (Whatman) and immediately loaded onto C18 SPE cartridges (Sep-Pak® Vac 3 cc tC18, Waters, Ireland) previously conditioned with 5 mL of HPLC water and methanol at a flow rate of 1 mL min⁻¹. Samples (500 mL) were loaded at a flow rate of 5 mL min⁻¹. After pre-concentration, the cartridges were completely dried in vacuo for 20 minutes
160 to avoid hydrolysis and kept frozen until analysis. Thereafter, cartridges were eluted with 4 mL
161 of methanol. Eluted samples were partially evaporated under a gentle nitrogen stream and
162 reconstituted in a final volume of 1 mL methanol. Samples were then analyzed by liquid
163 chromatography. The HPLC system comprised a binary HPLC Pump (Waters 1525), an auto
164 sampler (Waters 717 Plus) and a UV-detector (Waters 2487 Dual λ Abs. Detector). The HPLC
165 separation entailed use of a µm C₁₈ reverse-phase column (Sunfire 4.6x150 mm). For DIU
166 analysis, the injection volume was set at 20 µL and separation was performed using an isocratic
167 gradient of 45% methanol/55% water at 0.8 mL min⁻¹. The DIU peak was detected at 251 nm.
168 For TCS and Me-TCS analysis the injection volume was set at 50 µL and the flow rate was 1
169 mL min⁻¹ of 90% methanol with isocratic flow. The TCS and Me-TCS peaks were detected at
170 280 nm.

## 2.3 Biofilm structure and function

171 Several biofilm endpoints were measured in order to describe structural and functional responses
172 of autotrophic - and heterotrophic biofilm compartments of biofilms to toxicants pulses. In
173 particular the structure of the autotrophic community was investigated by measuring the
174 Chlorophyll-a density (as surrogate of autotrophic biomass) as well as the growth rate, mortality
175 and composition of diatom community. The structural response of heterotrophs was investigated
176 by counting the live and dead bacterial cells. This counting allowed the calculation of live/dead
177 ratio and therefore provide information about bacterial mortality. Scanning Electron Microscopy
178 (SEM) observations were useful to compare the structure of treated and non-treated biofilm at
179 the end of the experiment. The *in vivo* fluorescence measurements described the functional
180 response of phototrophic organisms (green algae, diatoms and cyanobacteria). The extracellular
181 enzymatic activities described the heterotrophic capacity to degrade organic matter. Finally, the
182 phosphorus uptake rate measurement was used to describe the capacity of the whole biofilm
183 community to remove phosphate from the water column.
184
to avoid hydrolysis and kept frozen until analysis. Thereafter, cartridges were eluted with 4 mL of methanol. Eluted samples were partially evaporated under a gentle nitrogen stream and reconstituted in a final volume of 1 mL methanol. Samples were then analyzed by liquid chromatography. The HPLC system comprised a binary HPLC Pump (Waters 1525), an auto sampler (Waters 717 Plus) and a UV-detector (Waters 2487 Dual λ Abs. Detector). The HPLC separation entailed use of a µm C₁₈ reverse-phase column (Sunfire 4.6x150 mm). For DIU analysis, the injection volume was set at 20 µL and separation was performed using an isocratic gradient of 45% methanol/55% water at 0.8 mL min⁻¹. The DIU peak was detected at 251 nm. For TCS and Me-TCS analysis the injection volume was set at 50 µL and the flow rate was 1 mL min⁻¹ of 90% methanol with isocratic flow. The TCS and Me-TCS peaks were detected at 280 nm.

## 2.3 Biofilm structure and function

Several biofilm endpoints were measured in order to describe structural and functional responses of autotrophic - and heterotrophic biofilm compartments of biofilms to toxicants pulses. In particular the structure of the autotrophic community was investigated by measuring the Chlorophyll-a density (as surrogate of autotrophic biomass) as well as the growth rate, mortality and composition of diatom community. The structural response of heterotrophs was investigated by counting the live and dead bacterial cells. This counting allowed the calculation of live/dead ratio and therefore provide information about bacterial mortality. Scanning Electron Microscopy (SEM) observations were useful to compare the structure of treated and non-treated biofilm at the end of the experiment. The *in vivo* fluorescence measurements described the functional response of phototrophic organisms (green algae, diatoms and cyanobacteria). The extracellular enzymatic activities described the heterotrophic capacity to degrade organic matter. Finally, the phosphorus uptake rate measurement was used to describe the capacity of the whole biofilm community to remove phosphate from the water column.
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**2.3.1 Chlorophyll-a density.** Glass tiles for chlorophyll analysis were collected from the

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mesocosms and stored in the dark at -20 °C until analysis. Chlorophyll-a in the biofilms was

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quantified after extraction from the glass substrata in 90% acetone in the dark at 4 °C for 12 h.

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Concentration was determined spectrophotometrically after filtration (GF/F Whatman) of the

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extract, following the procedure of Jeffrey & Humphrey (1975).

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**2.3.2 Diatom community structure and live/dead ratio.** Glass slides were scraped using

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polyethylene cell lifters (Corning Inc., NY), preserved with a drop of formalin solution, and then

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diluted with artificial stream water to a final volume of 5 mL. Samples were ultrasonicated for 7

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minutes (sonication bath Bransonic 220, Technofix) to separate the aggregated cells. A 125 µL

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aliquot of each sample were then pipetted onto a Nageotte counting chamber. The cells were

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counted in ten fields (1.25 µL each, 0.5 mm depth) by light microscopy (Nikon Eclipse 80i,

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Nikon Co., Tokyo, Japan) at 10× magnification. Data were recorded as the number of cells per

199

unit area of sampled substrate (number of cells per cm²), and the cells were classified as either

200

live (if they contained chloroplasts) or dead (if they were empty) (Morin at al., 2010b). Growth

201

rates (expressed as number of cell divisions per day) of the diatom community were calculated

202

from live diatom counts (Guillard, 1973). Diatom community composition was determined based

203

on permanent slides following ANSP protocols (Charles *et al.* 2002). Approximately 500

204

frustules were counted per slide at 1,000× magnification; they were identified using European

205

taxonomic literature (Krammer and Lange-Bertalot, 1986-1991) and recent nomenclature

206

updates.

207

**2.3.3 Bacterial density.** Live and dead bacteria were counted with epifluorescence microscopy

208

using the LIVE/DEAD® Bacteria Viability Kit L7012 (BacLight™, Molecular Probes,

209

Invitrogen). Biofilm samples were sonicated (< 1 min, sonication bath at 40 W and 40 kHz,
185

186

187 **2.3.1 Chlorophyll-a density.** Glass tiles for chlorophyll analysis were collected from the mesocosms and stored in the dark at -20 °C until analysis. Chlorophyll-a in the biofilms was quantified after extraction from the glass substrata in 90% acetone in the dark at 4 °C for 12 h.
188
189
190 Concentration was determined spectrophotometrically after filtration (GF/F Whatman) of the extract, following the procedure of Jeffrey & Humphrey (1975).
191

192 **2.3.2 Diatom community structure and live/dead ratio.** Glass slides were scraped using polyethylene cell lifters (Corning Inc., NY), preserved with a drop of formalin solution, and then diluted with artificial stream water to a final volume of 5 mL. Samples were ultrasonicated for 7 minutes (sonication bath Bransonic 220, Technofix) to separate the aggregated cells. A 125 µL aliquot of each sample were then pipetted onto a Nageotte counting chamber. The cells were counted in ten fields (1.25 µL each, 0.5 mm depth) by light microscopy (Nikon Eclipse 80i, Nikon Co., Tokyo, Japan) at 10× magnification. Data were recorded as the number of cells per unit area of sampled substrate (number of cells per cm²), and the cells were classified as either live (if they contained chloroplasts) or dead (if they were empty) (Morin at al., 2010b). Growth rates (expressed as number of cell divisions per day) of the diatom community were calculated from live diatom counts (Guillard, 1973). Diatom community composition was determined based on permanent slides following ANSP protocols (Charles *et al.* 2002). Approximately 500 frustules were counted per slide at 1,000× magnification; they were identified using European taxonomic literature (Krammer and Lange-Bertalot, 1986-1991) and recent nomenclature updates.
193
194
195
196
197
198
199
200
201
202
203
204
205
206

207 **2.3.3 Bacterial density.** Live and dead bacteria were counted with epifluorescence microscopy using the LIVE/DEAD® Bacteria Viability Kit L7012 (*BacLight™*, Molecular Probes, 208
209 Invitrogen). Biofilm samples were sonicated (< 1 min, sonication bath at 40 W and 40 kHz,
210 Selecta) and scraped (Nunc sterile silicone cell scraper) to obtain a biofilm suspension. Samples 211 were then diluted with pre-filtered sterilized water from the mesocosms, and 2 mL subsamples 212 were stained with 3 µL of a 1:1 mixture of SYTO 9 and propidium iodide for 15-30 minutes in 213 the dark. After incubation, samples were filtered through a 0.2 µm black polycarbonate filters 214 (Nuclepore, Whatman). Filters were then dried, placed on a slide with mounting oil (Molecular 215 Probes) and counted by epifluorescence microscopy (Nikon E600, 100× in immersion oil). Green 216 (live) and red (dead) bacterial cells were counted in 20 random fields per filter. 217

217 **2.3.4 SEM observation.** One glass tile from each mesocosm was collected at day 16 for SEM 218 observation. Samples were fixed immediately with 2.5% glutaraldehyde in 0.1 M cacodylate 219 buffer, pH 7.2–7.4. Samples were dehydrated in graded ethanol (65–100%) and dried at the 220 critical point of CO2. Finally, samples were sputter coated with gold, and then observed by SEM 221 (Zeiss DSM 960). 222

222 **2.3.5 In vivo fluorescence measurements.** Fluorescence emission from chlorophyll was 223 measured with a PhytoPAM (Pulse Amplitude Modulated) fluorometer (Heinz Walz GmbH), 224 which uses a set of LEDs at four wavelengths (470, 520, 645, and 665 nm). Five analytical 225 replicates (1 cm⁻² glass tiles) for each experimental replicate (mesocosm) were analyzed and 226 averaged (Serra *et al.*, 2009). The photosynthetic efficiency (Yₐff) of photosystem II (PSII) and 227 the photosynthetic capacity (Ymax) of PSII were also estimated (Ricart *et al.*, 2010). 228

228 **2.3.6 Extracellular enzymatic activity.** Leucine-aminopeptidase (EC3.4.11.1) and alkaline 229 phosphatase (EC 3.1.3.1-2) in the biofilms were quantified spectrofluorometrically using the 230 fluorescent-linked substrates L-leucine-4-methyl-7-coumarinylamide (Leu-AMC, Sigma-Aldrich) 231 and MUF-phosphate (MUF-P, Sigma-Aldrich), respectively. Colonized glass substrata were 232 collected from the mesocosms and placed in vials filled with 4 mL of pre-filtered water (0.2 µm 233 Nylon Membrane filters, Whatman). Samples were immediately incubated at saturating 234 conditions (0.3 mM, Romaní and Sabater, 1999) for 1 h in the dark in a shaking bath. Blanks and
210 Selecta) and scraped (Nunc sterile silicone cell scraper) to obtain a biofilm suspension. Samples 211 were then diluted with pre-filtered sterilized water from the mesocosms, and 2 mL subsamples 212 were stained with 3 µL of a 1:1 mixture of SYTO 9 and propidium iodide for 15-30 minutes in 213 the dark. After incubation, samples were filtered through a 0.2 µm black polycarbonate filters 214 (Nuclepore, Whatman). Filters were then dried, placed on a slide with mounting oil (Molecular 215 Probes) and counted by epifluorescence microscopy (Nikon E600, 100× in immersion oil). Green 216 (live) and red (dead) bacterial cells were counted in 20 random fields per filter. 217

217 **2.3.4 SEM observation.** One glass tile from each mesocosm was collected at day 16 for SEM 218 observation. Samples were fixed immediately with 2.5% glutaraldehyde in 0.1 M cacodylate 219 buffer, pH 7.2-7.4. Samples were dehydrated in graded ethanol (65-100%) and dried at the 220 critical point of CO2. Finally, samples were sputter coated with gold, and then observed by SEM 221 (Zeiss DSM 960). 222

222 **2.3.5 In vivo fluorescence measurements.** Fluorescence emission from chlorophyll was 223 measured with a PhytoPAM (Pulse Amplitude Modulated) fluorometer (Heinz Walz GmbH), 224 which uses a set of LEDs at four wavelengths (470, 520, 645, and 665 nm). Five analytical 225 replicates (1 cm⁻² glass tiles) for each experimental replicate (mesocosm) were analyzed and 226 averaged (Serra *et al.*, 2009). The photosynthetic efficiency (Yₐff) of photosystem II (PSII) and 227 the photosynthetic capacity (Ymax) of PSII were also estimated (Ricart *et al.*, 2010). 228

228 **2.3.6 Extracellular enzymatic activity.** Leucine-aminopeptidase (EC3.4.11.1) and alkaline 229 phosphatase (EC 3.1.3.1-2) in the biofilms were quantified spectrofluorometrically using the 230 fluorescent-linked substrates L-leucine-4-methyl-7-coumarinylamide (Leu-AMC, Sigma-Aldrich) 231 and MUF-phosphate (MUF-P, Sigma-Aldrich), respectively. Colonized glass substrata were 232 collected from the mesocosms and placed in vials filled with 4 mL of pre-filtered water (0.2 µm 233 Nylon Membrane filters, Whatman). Samples were immediately incubated at saturating 234 conditions (0.3 mM, Romaní and Sabater, 1999) for 1 h in the dark in a shaking bath. Blanks and
control samples were used to correct non-enzymatic hydrolysis of the substrate and of fluorescent substances in the solution, respectively. Blanks, control samples, and AMC and MUF standards (0–100 µM) were also incubated in the shaking bath. Following incubation, 4 mL of glycine buffer (pH 10.4) solution was added (1:1, v/v), and fluorescence was measured at 364 and 445 nm (excitation & emission) for AMC and at 365 and 455 nm (excitation & emission) for MUF (Kontron, SFM25). The calculated AMC and MUF concentrations were then standardized for glass surface and incubation time; therefore, the activity values are expressed as nanomoles (of AMC or MUF) per cm² of biofilm surface area per hour. To determine the specific activity per live cell, the activities per unit area were further normalized per the number of live bacteria (Ricart *et al.*, 2009).

**2.3.7 Phosphorus uptake.** Soluble reactive phosphorus (SRP) uptake (P-uptake) of the biofilms was calculated by measuring the decay of SRP following a spike. Background SRP concentrations were analyzed at each sampling date. SRP concentration was increased by adding a spike of 10 mM Na₂PO₄, which increased basal concentrations by roughly four to eight times. Samples for SRP concentration (10 mL) were taken five times between 1 and 20 minutes post-spike in each replicate. The P-uptake was calculated as the mass of phosphorus removed from the water column per unit area per time (µg P cm⁻² h⁻¹). Abiotic controls showed that SRP had not decayed during the experiment.

## 2.4 Statistical tests

Differences in the biofilm endpoints were tested daily using one-way analysis of variance (ANOVA), in which treatment (DIU or TCS) was set as the fixed factor. Effects were analyzed *post hoc* with Tukey's b test. Statistical significance was set at p = 0.05. Analysis was performed using SPSS Version 15.0. Growth rates were calculated from the slope of the linear portion of a
control samples were used to correct non-enzymatic hydrolysis of the substrate and of fluorescent substances in the solution, respectively. Blanks, control samples, and AMC and MUF standards (0–100 µM) were also incubated in the shaking bath. Following incubation, 4 mL of glycine buffer (pH 10.4) solution was added (1:1, v/v), and fluorescence was measured at 364 and 445 nm (excitation & emission) for AMC and at 365 and 455 nm (excitation & emission) for MUF (Kontron, SFM25). The calculated AMC and MUF concentrations were then standardized for glass surface and incubation time; therefore, the activity values are expressed as nanomoles (of AMC or MUF) per cm² of biofilm surface area per hour. To determine the specific activity per live cell, the activities per unit area were further normalized per the number of live bacteria (Ricart *et al.*, 2009).

**2.3.7 Phosphorus uptake.** Soluble reactive phosphorus (SRP) uptake (P-uptake) of the biofilms was calculated by measuring the decay of SRP following a spike. Background SRP concentrations were analyzed at each sampling date. SRP concentration was increased by adding a spike of 10 mM Na₂PO₄, which increased basal concentrations by roughly four to eight times. Samples for SRP concentration (10 mL) were taken five times between 1 and 20 minutes post-spike in each replicate. The P-uptake was calculated as the mass of phosphorus removed from the water column per unit area per time (µg P cm⁻² h⁻¹). Abiotic controls showed that SRP had not decayed during the experiment.

## 2.4 Statistical tests

Differences in the biofilm endpoints were tested daily using one-way analysis of variance (ANOVA), in which treatment (DIU or TCS) was set as the fixed factor. Effects were analyzed *post hoc* with Tukey's b test. Statistical significance was set at p = 0.05. Analysis was performed using SPSS Version 15.0. Growth rates were calculated from the slope of the linear portion of a
curve showing the log of the cell number as a function of time, as described by Morin *et al.*
(2008), and tested with ANOVA.

## 3. RESULTS

### 3.1 Physical and chemical conditions in the mesocosms

Conductivity, pH and dissolved oxygen in the mesocosms remained steady during the experiment: their mean values were 143.4 (± 34.1) µS cm⁻¹, 7.9 (± 0.5), and 9.2 (± 0.2) mg L⁻¹ (n= 48), respectively. SRP concentration ranged from 16.9 (± 2.1) to 4.5 (±1.9) µg L⁻¹ between water replacements. pH was monitored between water changes and ranged from 7.45 (± 0.02) to 7.62 (±0.04) (n = 8), never reaching the TCS pKa value of 8.1.

During the 48 hours of toxicant exposure, the DIU-contaminated mesocosms had 13.4 ± 1.3 µg DIU L⁻¹ (n = 4), and the TCS-contaminated mesocosms, 60.8 ± 30.1 µg TCS L⁻¹ (n = 4).

The level of Me-TCS in the TCS-contaminated mesocosms was consistently below the detection limit. The water of control mesocoms resulted in no trace of both DIU and TCS (n = 8).

### 3.2 Biofilm microbial biomass, diatom community composition, and SEM observations

Before toxicant exposure, the biofilms had chlorophyll-a concentration of 3.04 ± 0.72 µg cm⁻² and a mean bacterial density of 3.06 ± 1.14 x 10⁷ cells cm⁻². The diatom community was dominated by *Achnanthidium minutissimum* (Kützing), and also contained *Achnanthidium biasolettianum* (Grunow), *Ulnaria ulna* (Nitzsch) and *Gomphonema*. Live diatom density was 11.3 ± 3.7 x 10⁴ cells cm⁻², and the diatom live/dead ratio was 10.9 ± 0.9. The bacteria live/dead ratio was 1.48 ± 0.58.

Chlorophyll-a density increased in all the treatments from day 0 to day 16 (Fig. 1a); however, that of the DIU-treated biofilms was not significantly different than that of the control biofilms. Chlorophyll-a density in TCS-treated biofilms decreased relative to that of the control by day 2, and subsequently increased significantly (p = 0.026 and p = 0.007, for days 9 and 16, respectively) (Fig. 1a).
curve showing the log of the cell number as a function of time, as described by Morin *et al.*
(2008), and tested with ANOVA.

## 3. RESULTS

### 3.1 Physical and chemical conditions in the mesocosms

Conductivity, pH and dissolved oxygen in the mesocosms remained steady during the experiment: their mean values were 143.4 (± 34.1) µS cm⁻¹, 7.9 (± 0.5), and 9.2 (± 0.2) mg L⁻¹ (n= 48), respectively. SRP concentration ranged from 16.9 (± 2.1) to 4.5 (±1.9) µg L⁻¹ between water replacements. pH was monitored between water changes and ranged from 7.45 (± 0.02) to 7.62 (±0.04) (n = 8), never reaching the TCS pKa value of 8.1.

During the 48 hours of toxicant exposure, the DIU-contaminated mesocosms had 13.4 ± 1.3 µg DIU L⁻¹ (n = 4), and the TCS-contaminated mesocosms, 60.8 ± 30.1 µg TCS L⁻¹ (n = 4). The level of Me-TCS in the TCS-contaminated mesocosms was consistently below the detection limit. The water of control mesocoms resulted in no trace of both DIU and TCS (n = 8).

### 3.2 Biofilm microbial biomass, diatom community composition, and SEM observations

Before toxicant exposure, the biofilms had chlorophyll-a concentration of 3.04 ± 0.72 µg cm⁻² and a mean bacterial density of 3.06 ± 1.14 x 10⁷ cells cm⁻². The diatom community was dominated by *Achnanthidium minutissimum* (Kützing), and also contained *Achnanthidium biasolettianum* (Grunow), *Ulnaria ulna* (Nitzsch) and *Gomphonema*. Live diatom density was 11.3 ± 3.7 x 10⁴ cells cm⁻², and the diatom live/dead ratio was 10.9 ± 0.9. The bacteria live/dead ratio was 1.48 ± 0.58.

Chlorophyll-a density increased in all the treatments from day 0 to day 16 (Fig. 1a); however, that of the DIU-treated biofilms was not significantly different than that of the control biofilms. Chlorophyll-a density in TCS-treated biofilms decreased relative to that of the control by day 2, and subsequently increased significantly (p = 0.026 and p = 0.007, for days 9 and 16, respectively) (Fig. 1a).
Live diatom density increased exponentially (Fig. 2) from day 0 to day 16 in all treatments. Diatoms growth rates (Fig. 2) were higher (0.14 divisions day⁻¹) in DIU-treated biofilms than in either control or TCS-treated biofilms (Fig. 2; *p* = 0.001). Diatom composition did not significantly change in either DIU or control biofilms, but *Achnanthidium minutissimum* became dominant in the TCS-treated biofilms. The diatom live/dead ratio significantly decreased in DIU-treated biofilms starting from day 2 and remained lower than that of the control biofilms until the end of experiment (Fig. 1b). Diatoms of TCS-treated biofilms responded late to exposure (Fig. 1b): the live/dead ratio significantly decreased by day 9 (7.4 ± 0.7; 69% of the control value; *p* = 0.001), recovering moderately by day 16 (Fig. 1b).

The SEM showed that the filamentous green alga *Spirogyra* sp. thrived with unharmed filaments both in control and DIU-contaminated mesocosms (Fig. 3, day 16). However, the *Spirogyra* filaments were less abundant and visually damaged, hosting abundant epibionts (*Achnanthidium* species in Fig. 3b and c) in TCS-treated mesocosms. The green algae contribution to the total chlorophyll *a* content of the TCS-treated samples was less than 0.5% (data derived from fluorescence, not shown).

DIU exposure did not affect bacterial mortality. However, TCS did cause a significant but temporary surge in bacterial mortality (145% of the control value; *p* < 0.001; Fig. 1c).

## 3.3 Biofilm function

The results from photosynthetic capacity are included in Figure 4 while results from extracellularar enzymes are reported in this section. The extracellular enzymatic activity and the physiological parameters of the biofilms were similar among mesocosms before toxicant exposure. Leucine-aminopeptidase activity was 419 ± 78 nmol AMC cm⁻² h⁻¹ and alkaline phosphatase activity was 136 ± 21 nmol MUF cm⁻² h⁻¹. Photosynthetic capacity was 0.47 ± 0.06 photon yield and the photosynthetic efficiency was 0.4 ± 0.05 photon yield. The P-uptake before exposure was 2.1 ± 0.3 µg P cm⁻² h⁻¹.
Live diatom density increased exponentially (Fig. 2) from day 0 to day 16 in all treatments. Diatoms growth rates (Fig. 2) were higher (0.14 divisions day⁻¹) in DIU-treated biofilms than in either control or TCS-treated biofilms (Fig. 2; *p* = 0.001). Diatom composition did not significantly change in either DIU or control biofilms, but *Achnanthidium minutissimum* became dominant in the TCS-treated biofilms. The diatom live/dead ratio significantly decreased in DIU-treated biofilms starting from day 2 and remained lower than that of the control biofilms until the end of experiment (Fig. 1b). Diatoms of TCS-treated biofilms responded late to exposure (Fig. 1b): the live/dead ratio significantly decreased by day 9 (7.4 ± 0.7; 69% of the control value; *p* = 0.001), recovering moderately by day 16 (Fig. 1b).

The SEM showed that the filamentous green alga *Spirogyra* sp. thrived with unharmed filaments both in control and DIU-contaminated mesocosms (Fig. 3, day 16). However, the *Spirogyra* filaments were less abundant and visually damaged, hosting abundant epibionts (*Achnanthidium* species in Fig. 3b and c) in TCS-treated mesocosms. The green algae contribution to the total chlorophyll *a* content of the TCS-treated samples was less than 0.5% (data derived from fluorescence, not shown).

DIU exposure did not affect bacterial mortality. However, TCS did cause a significant but temporary surge in bacterial mortality (145% of the control value; *p* < 0.001; Fig. 1c).

## 3.3 Biofilm function

The results from photosynthetic capacity are included in Figure 4 while results from extracellular enzymes are reported in this section. The extracellular enzymatic activity and the physiological parameters of the biofilms were similar among mesocosms before toxicant exposure. Leucine-aminopeptidase activity was 419 ± 78 nmol AMC cm⁻² h⁻¹ and alkaline phosphatase activity was 136 ± 21 nmol MUF cm⁻² h⁻¹. Photosynthetic capacity was 0.47 ± 0.06 photon yield and the photosynthetic efficiency was 0.4 ± 0.05 photon yield. The P-uptake before exposure was 2.1 ± 0.3 µg P cm⁻² h⁻¹.
Extracellular enzyme activity did not differ among the different biofilms and generally increased with time. Phosphatase activity increased up to 266 ± 28 nmol MUF cm⁻² h⁻¹ (day 9), and leucine-aminopeptidase increased up to 544 ± 58 nmol MUF cm⁻² h⁻¹ (day 9). These temporary changes were not associated to the effects of DIU or TCS. In contrast, DIU had enhanced specific extracellular enzyme activity per cell by day 2: the DIU-treated biofilms exhibited an increase in specific alkaline phosphatase and leucine aminopeptidase activity per cell, up to 250 % of the values in the control biofilms (data not shown). Both specific activities returned to normal levels by 1 week post-exposure (day 9).

After 48 hours treatment, DIU significantly affected both photosynthetic efficiency (p < 0.001) and capacity (p = 0.004), whereas TSC did not significantly affect either one (Fig. 4). However, in the DIU-treated biofilms, both photosynthetic parameters returned to normal levels by 1 week post-exposure.

P-uptake was not affected in the DIU-treated biofilms, but decreased significantly in the TCS-treated biofilms (Fig. 5): at day 2 its value was 1.1 ± 0.3 µg P cm⁻² h⁻¹ (29.3 % of the control value; p < 0.001). The TCS-induced effects persisted until 1 week post-exposure: at day 9 P-uptake was 2.1 ± 0.1 µg P cm⁻² h⁻¹ (77.4 % of the control value; p = 0.001), although by day 16, it had returned to normal levels (Fig. 5).

## 4. DISCUSSION

Several studies have investigated the effects of DIU (Pesce et al., 2006; Ricart et al., 2009; Tlili et al., 2008) and TCS on river biofilms (Franz et al, 2008; Lawrence et al., 2009; Ricart et al., 2010). Nevertheless, these studies mainly used dose-response designs, and exposure of biofilms to the toxicants, to describe the effects of chronic contamination. The results reported in the present work show that short pulses of either compound also affect natural biofilm
309 Extracellular enzyme activity did not differ among the different biofilms and generally
310 increased with time. Phosphatase activity increased up to 266 ± 28 nmol MUF cm⁻² h⁻¹ (day 9),
311 and leucine-aminopeptidase increased up to 544 ± 58 nmol MUF cm⁻² h⁻¹ (day 9). These
temporary changes were not associated to the effects of DIU or TCS. In contrast, DIU had
312 enhanced specific extracellular enzyme activity per cell by day 2: the DIU-treated biofilms
313 exhibited an increase in specific alkaline phosphatase and leucine aminopeptidase activity per
cell, up to 250 % of the values in the control biofilms (data not shown). Both specific activities
315 returned to normal levels by 1 week post-exposure (day 9).
316

317 After 48 hours treatment, DIU significantly affected both photosynthetic efficiency (p <
318 0.001) and capacity (p = 0.004), whereas TSC did not significantly affect either one (Fig. 4).
319 However, in the DIU-treated biofilms, both photosynthetic parameters returned to normal levels
320 by 1 week post-exposure.
321

321 P-uptake was not affected in the DIU-treated biofilms, but decreased significantly in the
322 TCS-treated biofilms (Fig. 5): at day 2 its value was 1.1 ± 0.3 µg P cm⁻² h⁻¹ (29.3 % of the control
323 value; p < 0.001). The TCS-induced effects persisted until 1 week post-exposure: at day 9 P-
324 uptake was 2.1 ± 0.1 µg P cm⁻² h⁻¹ (77.4 % of the control value; p = 0.001), although by day 16, it
325 had returned to normal levels (Fig. 5).
326

326

327 **4. DISCUSSION**

328 Several studies have investigated the effects of DIU (Pesce *et al.*, 2006; Ricart *et al.*, 2009; Tlili *et al.*, 2008) and TCS on river biofilms (Franz et al, 2008; Lawrence *et al.*, 2009; Ricart *et al.*, 2010). Nevertheless, these studies mainly used dose-response designs, and exposure of biofilms to the toxicants, to describe the effects of chronic contamination. The results reported in the present work show that short pulses of either compound also affect natural biofilm
329
330
331
332
communities, and that post-pulse behavior depends on the toxicant used and on the endpoint considered. Due to the different target of the two toxicants and the interactions between autotrophs and heterotrophs within the biofilm, direct and indirect effects are highlighted. Moreover, the study of the post-pulse behavior shows the timing of these effects (either rapid or delayed) as well as potential recovery (return to values not significantly different respect to the control). The multi-biomarker approach (Boninneau et al., 2010) employed in this study enabled description of direct and indirect effects, as well as their recovery, associated with short pulses of either toxicant. Short-term pulses can be considered as transitory perturbations, which can generate responses in the structure and function of fluvial biofilms. Whether the responses after these short-term perturbations are immediate or delayed in time, responses depend on the organisms directly or indirectly targeted by the stressor, as well as on the mechanism associated to the measured parameter. Once a significant response occurred (i.e. increase or decrease of some activity; increase or decrease of mortality; shift in community composition etc.), the effects can persist in time, or instead recover to the original status. We considered that parameters recovered when values after the disturbance were close to values in the control after the disturbance. However, the significance of perturbations and the recovery at the ecosystem scale is a more complex subject that the one being dealt in the paper at a mesocosms scale, and therefore cannot be directly extrapolated to real systems. The short pulses of DIU inhibited photosynthetic efficiency and capacity and increased diatom mortality. These effects are related to its inhibition of photosynthesis via blockage of electron transport in photosystem II (Van Rensen, 1989). Several studies have confirmed this effect on natural epipelagic and epilithic biofilm communities (Legrand et al., 2006; López-Doval et al., 2010). However, the short-term pulses had transient functional effects on the autotrophs and the photosynthetic parameters had rapidly recovered (returned to control values) by 1 week post-exposure. On the other hand, recovery of functional parameters may hide specific effects on community (i.e. species replacement, composition shift).
communities, and that post-pulse behavior depends on the toxicant used and on the endpoint considered. Due to the different target of the two toxicants and the interactions between autotrophs and heterotrophs within the biofilm, direct and indirect effects are highlighted. Moreover, the study of the post-pulse behavior shows the timing of these effects (either rapid or delayed) as well as potential recovery (return to values not significantly different respect to the control). The multi-biomarker approach (Boninneau et al., 2010) employed in this study enabled description of direct and indirect effects, as well as their recovery, associated with short pulses of either toxicant. Short-term pulses can be considered as transitory perturbations, which can generate responses in the structure and function of fluvial biofilms. Whether the responses after these short-term perturbations are immediate or delayed in time, responses depend on the organisms directly or indirectly targeted by the stressor, as well as on the mechanism associated to the measured parameter. Once a significant response occurred (i.e. increase or decrease of some activity; increase or decrease of mortality; shift in community composition etc.), the effects can persist in time, or instead recover to the original status. We considered that parameters recovered when values after the disturbance were close to values in the control after the disturbance. However, the significance of perturbations and the recovery at the ecosystem scale is a more complex subject that the one being dealt in the paper at a mesocosms scale, and therefore cannot be directly extrapolated to real systems. The short pulses of DIU inhibited photosynthetic efficiency and capacity and increased diatom mortality. These effects are related to its inhibition of photosynthesis via blockage of electron transport in photosystem II (Van Rensen, 1989). Several studies have confirmed this effect on natural epipelagic and epilithic biofilm communities (Legrand et al., 2006; López-Doval et al., 2010). However, the short-term pulses had transient functional effects on the autotrophs and the photosynthetic parameters had rapidly recovered (returned to control values) by 1 week post-exposure. On the other hand, recovery of functional parameters may hide specific effects on community (i.e. species replacement, composition shift).
Ricart et al (2009) evidenced how chronic DIU exposure induced shift in diatoms community composition and decrease in diatoms biovolume. Nevertheless the same study also evidenced that photosynthetic parameters did not recover despite the shift of community composition. The recovery of photosynthetic parameters in our study occurred despite the low resistance of diatoms and their slow recovery after 48 hours of short pulses (10 µg L⁻¹) of DIU. The significant increase in diatom growth rate (Fig. 2) and the absence of shift in community composition after the DIU pulses might indicate that the diatoms recovery was occurring at the end of the experiment. These results could be explained by the short time of exposure leading to transient direct effects on photosynthesis and on diatoms viability. A relevant side-effect was that algal biomass (chlorophyll-a) was moderately enhanced by DIU exposure. Algal biomass (chlorophyll) was moderately enhanced by DIU exposure. Other authors have also observed this increase in chlorophyll density and have related it to the interruption of electron flow in PSII provoked by DIU (Ricart et al., 2009; Tlili et al., 2008), as well as to the induction of shade-type chloroplasts with a higher concentration of photosynthetic pigments (Chesworth et al., 2004). On the other hand any significant effect of DIU on bacterial viability has been observed. In other studies, DIU did indirectly affect bacteria mortality and extracellular enzyme activity, but these studies involved long-term exposure (Ricart et al., 2009). Moreover, chronic exposure has been demonstrated to induce shift in bacterial community composition of biofilms in the case of DIU (Pesce et al 2006; Tlili et al., 2008)

The moderate effects on periphyton structure and function after the DIU pulses showed in this study are consistent with the three-stage model proposed by Mølander and Blanck (1992). This model joins various effects of DIU on periphyton structure and function. In the first stage no long-term effects can be detected in spite of short-term effects, such as inhibition of photosynthesis. The second stage would be characterized by slight long-term effects such as the increase of chlorophyll-a. The final stage would occur when the sensitive species would be
Ricart et al (2009) evidenced how chronic DIU exposure induced shift in diatoms community composition and decrease in diatoms biovolume. Nevertheless the same study also evidenced that photosynthetic parameters did not recover despite the shift of community composition. The recovery of photosynthetic parameters in our study occurred despite the low resistance of diatoms and their slow recovery after 48 hours of short pulses (10 µg L⁻¹) of DIU. The significant increase in diatom growth rate (Fig. 2) and the absence of shift in community composition after the DIU pulses might indicate that the diatoms recovery was occurring at the end of the experiment. These results could be explained by the short time of exposure leading to transient direct effects on photosynthesis and on diatoms viability. A relevant side-effect was that algal biomass (chlorophyll-a) was moderately enhanced by DIU exposure. Algal biomass (chlorophyll) was moderately enhanced by DIU exposure. Other authors have also observed this increase in chlorophyll density and have related it to the interruption of electron flow in PSII provoked by DIU (Ricart et al., 2009; Tlili et al., 2008), as well as to the induction of shade-type chloroplasts with a higher concentration of photosynthetic pigments (Chesworth et al., 2004). On the other hand any significant effect of DIU on bacterial viability has been observed. In other studies, DIU did indirectly affect bacteria mortality and extracellular enzyme activity, but these studies involved long-term exposure (Ricart et al., 2009). Moreover, chronic exposure has been demonstrated to induce shift in bacterial community composition of biofilms in the case of DIU (Pesce et al 2006; Tlili et al., 2008)

The moderate effects on periphyton structure and function after the DIU pulses showed in this study are consistent with the three-stage model proposed by Mølander and Blanck (1992). This model joins various effects of DIU on periphyton structure and function. In the first stage no long-term effects can be detected in spite of short-term effects, such as inhibition of photosynthesis. The second stage would be characterized by slight long-term effects such as the increase of chlorophyll-a. The final stage would occur when the sensitive species would be
383 eliminated resulting in restructured community and increased community tolerance. Achieving
384 this last stage should imply that the diuron stress would be sufficiently severe to cause cell
385 mortality. The recovery of photosynthetic parameters and the increase of chlorophyll-a density
386 evidenced in this study suggest that 48 hours exposure to 13µg L⁻¹ of DIU can be considered a
387 threshold between first and second stage impact (Mølander and Blanck, 1992) in biofilm
388 communities.

389 In contrast to DIU, the mode of action of the bactericide TCS leads to a strong direct effect on
390 bacterial viability in the biofilm. TCS might be inhibiting fatty acid synthesis and bacterial
391 growth (Escalada *et al.*, 2005). However one week after the end of exposure, live/dead bacteria
392 ratio values were similar than controls indicating a recovery of the bacterial community.
393 Nevertheless, considering the biology and the short life cycle of bacteria, selection of resistant
394 species and consequent shift in community composition could occur, although this would be most
395 probably in response to longer exposure time. However, to our knowledge, no experimental data
396 are available about chronic contamination effects of TCS on biofilm bacteria.

397 Although TCS has chiefly been described as a bactericide (and was like this in the present
398 study, where the ratio of live/dead bacteria indeed decreased), it also significantly affected
399 autotrophs (non-target organisms). Diatom mortality increased for 1 week post-exposure (day 9),
400 and TCS exposure retarded development of the diatom community relative to the control. These
401 results may reflect a delayed direct effect of TCS on diatoms or an indirect effect of bacteria
402 mortality as a result of the tight interaction between these two biofilms components. The difficulty
403 of growing axenic cultures of benthic diatoms (Bruckner and Kroth, 2009) demonstrates that
404 diatoms require bacteria for proper development (e.g. bacteria vitamin production for algae; Croft
405 *et al.*, 2005). The delayed effect on diatoms (day 9) could have been a late indirect response to
406 the increase in bacterial mortality that had occurred on day 2. This scenario is corroborated by the
407 fact that the diatoms recovered within 1 week after the bacteria had recovered. Chlorophyll
383 eliminated resulting in restructured community and increased community tolerance. Achieving
384 this last stage should imply that the diuron stress would be sufficiently severe to cause cell
385 mortality. The recovery of photosynthetic parameters and the increase of chlorophyll-a density
386 evidenced in this study suggest that 48 hours exposure to 13µg L⁻¹ of DIU can be considered a
387 threshold between first and second stage impact (Mølander and Blanck, 1992) in biofilm
388 communities.

389 In contrast to DIU, the mode of action of the bactericide TCS leads to a strong direct effect on
390 bacterial viability in the biofilm. TCS might be inhibiting fatty acid synthesis and bacterial
391 growth (Escalada *et al.*, 2005). However one week after the end of exposure, live/dead bacteria
392 ratio values were similar than controls indicating a recovery of the bacterial community.
393 Nevertheless, considering the biology and the short life cycle of bacteria, selection of resistant
394 species and consequent shift in community composition could occur, although this would be most
395 probably in response to longer exposure time. However, to our knowledge, no experimental data
396 are available about chronic contamination effects of TCS on biofilm bacteria.

397 Although TCS has chiefly been described as a bactericide (and was like this in the present
398 study, where the ratio of live/dead bacteria indeed decreased), it also significantly affected
399 autotrophs (non-target organisms). Diatom mortality increased for 1 week post-exposure (day 9),
400 and TCS exposure retarded development of the diatom community relative to the control. These
401 results may reflect a delayed direct effect of TCS on diatoms or an indirect effect of bacteria
402 mortality as a result of the tight interaction between these two biofilms components. The difficulty
403 of growing axenic cultures of benthic diatoms (Bruckner and Kroth, 2009) demonstrates that
404 diatoms require bacteria for proper development (e.g. bacteria vitamin production for algae; Croft
405 *et al.*, 2005). The delayed effect on diatoms (day 9) could have been a late indirect response to
406 the increase in bacterial mortality that had occurred on day 2. This scenario is corroborated by the
407 fact that the diatoms recovered within 1 week after the bacteria had recovered. Chlorophyll
concentration had decreased after 48 hours of TCS exposure, but eventually returned to normal levels. The negative effect of TCS on chlorophyll-a, described elsewhere (White *et al.*, 2005), has been associated to modifications of biofilm architecture (Lawrence *et al.*, 2009). In the present work, SEM images (Fig. 3) showed that TCS had damaged *Spirogyra* sp. filaments and reduced chlorophyll density. The effects of TCS on the cell walls of *Spirogyra* sp. can be related to its blocking of fatty acid synthesis. This has been described in bacteria (McMurry *et al.*, 1998), and has been reported to compromise permeability-barrier functions (Phan and Marquis, 2006) and to destabilize cell membranes (Villaláin *et al.*, 2001). Although no specific mode-of-action for TCS has yet been established for algae, in some aspects this contaminant could affect algae similarly to the way it affects bacteria (Lawrence *et al.*, 2009; Ricart *et al.*, 2010; Morin *et al.*, 2010b). Moreover, some studies have described that algae are more sensitive to TCS than are bacteria (Tatarakazo *et al.*, 2004).

Neither DIU nor TCS affected the extracellular activity of either phosphatase or leucine-aminopeptidase in any of the biofilms. This indicates that the toxicant pulses did not compromise the ability of the biofilms to process organic matter (proteins) or organic phosphorus, despite the increased microbial mortality. Thus, one could infer that the biofilm maintained certain major functions even when its constituent organisms were directly or indirectly affected. This could be explained by either a change in the bacterial community or by a relatively higher specific activity per live cell (Francoeur and Wetzel, 2003). In fact, normalization of extracellular activities per live bacteria cell revealed that each toxicant had a distinct effect. In the TCS-treated biofilms, specific activity per cell was not affected, suggesting that an alternative mechanism dictates extracellular enzymatic activity in the biofilms. In contrast, the DIU-treated biofilms exhibited significant increases in both specific phosphatase and specific peptidase activity per live cell by day 2. Ricart *et al.* (2009) reported a similar indirect effect of DIU on the metabolism of live bacteria in the long-term. They concluded that the increase in extracellular leucine-
concentration had decreased after 48 hours of TCS exposure, but eventually returned to normal levels. The negative effect of TCS on chlorophyll-a, described elsewhere (White *et al.*, 2005), has been associated to modifications of biofilm architecture (Lawrence *et al.*, 2009). In the present work, SEM images (Fig. 3) showed that TCS had damaged *Spirogyra* sp. filaments and reduced chlorophyll density. The effects of TCS on the cell walls of *Spirogyra* sp. can be related to its blocking of fatty acid synthesis. This has been described in bacteria (McMurry *et al.*, 1998), and has been reported to compromise permeability-barrier functions (Phan and Marquis, 2006) and to destabilize cell membranes (Villaláin *et al.*, 2001). Although no specific mode-of-action for TCS has yet been established for algae, in some aspects this contaminant could affect algae similarly to the way it affects bacteria (Lawrence *et al.*, 2009; Ricart *et al.*, 2010; Morin *et al.*, 2010b). Moreover, some studies have described that algae are more sensitive to TCS than are bacteria (Tatarakazo *et al.*, 2004).

Neither DIU nor TCS affected the extracellular activity of either phosphatase or leucine-aminopeptidase in any of the biofilms. This indicates that the toxicant pulses did not compromise the ability of the biofilms to process organic matter (proteins) or organic phosphorus, despite the increased microbial mortality. Thus, one could infer that the biofilm maintained certain major functions even when its constituent organisms were directly or indirectly affected. This could be explained by either a change in the bacterial community or by a relatively higher specific activity per live cell (Francoeur and Wetzel, 2003). In fact, normalization of extracellular activities per live bacteria cell revealed that each toxicant had a distinct effect. In the TCS-treated biofilms, specific activity per cell was not affected, suggesting that an alternative mechanism dictates extracellular enzymatic activity in the biofilms. In contrast, the DIU-treated biofilms exhibited significant increases in both specific phosphatase and specific peptidase activity per live cell by day 2. Ricart *et al.* (2009) reported a similar indirect effect of DIU on the metabolism of live bacteria in the long-term. They concluded that the increase in extracellular leucine-
433

aminopeptidase activity per live cell was a response to the release of proteinaceous material from
DIU-induced lysis of algae cells.

434

435

DIU did not affect P-uptake, whereas TCS did. The bacterial death caused by TCS, together
with its delayed indirect effect on diatoms and its toxicity to Spirogyra sp., could have caused the
reduction in P-uptake. The fact that damage to heterotrophs and autotrophs generally had
negative consequences for P-uptake highlights the utility of this endpoint as a descriptor of total
biofilm function. Given that biofilms are the most important compartment in the biotic removal
of inorganic dissolved nutrients from water columns (Sabater et al., 2007), TCS-induced loss in
P-uptake is a clear threat to this ecological service of river ecosystems, which purifies water.

436

437

438

439

440

441

442

The results obtained with the large set of biomarkers confirm the central hypothesis of this
work: that direct effects on target organisms would occur earlier and are recovered in a short
period (especially if only physiological mechanisms were affected). The strong resistance of
bacteria to DIU, and the rapid recovery of photosynthetic parameters following exposure to DIU,
are consistent with the toxicant's specific mode-of-action, and support the aforementioned
hypothesis. Nevertheless, slow recovery of diatoms suggests that a combination of direct and
indirect effects could be at play. The hypothesis on delayed effects from indirect interactions was
confirmed by the results from the TCS exposure experiments. These effects on diatom mortality
appeared 1 week post-exposure and delayed biofilm recovery. Biofilms were more resistant and
resilient after the DIU pulses than after the TCS pulses. The fact that DIU did not affect P-uptake,
and that TCS did, agrees with the more complex behavior of direct and indirect effects associated
to this community function.

443

444

445

446

447

448

449

450

451

452

453

454

To conclude, the present study has confirmed the existence of direct effects of DIU and of
TCS on specific biofilm components, as well as indirect effects of each toxicant due to ecological
interactions within biofilms, as consequences of short-term pulses. Biofilms have shown their

455

456
433

aminopeptidase activity per live cell was a response to the release of proteinaceous material from
DIU-induced lysis of algae cells.

434

435

DIU did not affect P-uptake, whereas TCS did. The bacterial death caused by TCS, together
with its delayed indirect effect on diatoms and its toxicity to Spirogyra sp., could have caused the
reduction in P-uptake. The fact that damage to heterotrophs and autotrophs generally had
negative consequences for P-uptake highlights the utility of this endpoint as a descriptor of total
biofilm function. Given that biofilms are the most important compartment in the biotic removal
of inorganic dissolved nutrients from water columns (Sabater et al., 2007), TCS-induced loss in
P-uptake is a clear threat to this ecological service of river ecosystems, which purifies water.

436

437

438

439

440

441

442

The results obtained with the large set of biomarkers confirm the central hypothesis of this
work: that direct effects on target organisms would occur earlier and are recovered in a short
period (especially if only physiological mechanisms were affected). The strong resistance of
bacteria to DIU, and the rapid recovery of photosynthetic parameters following exposure to DIU,
are consistent with the toxicant's specific mode-of-action, and support the aforementioned
hypothesis. Nevertheless, slow recovery of diatoms suggests that a combination of direct and
indirect effects could be at play. The hypothesis on delayed effects from indirect interactions was
confirmed by the results from the TCS exposure experiments. These effects on diatom mortality
appeared 1 week post-exposure and delayed biofilm recovery. Biofilms were more resistant and
resilient after the DIU pulses than after the TCS pulses. The fact that DIU did not affect P-uptake,
and that TCS did, agrees with the more complex behavior of direct and indirect effects associated
to this community function.

443

444

445

446

447

448

449

450

451

452

453

454

To conclude, the present study has confirmed the existence of direct effects of DIU and of
TCS on specific biofilm components, as well as indirect effects of each toxicant due to ecological
interactions within biofilms, as consequences of short-term pulses. Biofilms have shown their

455

456
457 recovery capacity (by 2 weeks post-exposure they had recovered from nearly all the effects), but
458 also that even very short pulses of toxicants can have relevant consequences for biofilm structure
459 and function. It is reasonable to establish that short pulses can be seen as initial phases for effects
460 on biofilms, that longer toxicant pulses could imply more persistent effects and finally that
461 chronic concentrations of these toxicants could represent the most severe threat to biofilm
diversity and function.
462

463

464

465

466 Acknowledgements

467 This study was supported by the European Commission projects MODELKEY (Project 511237-2
468 GOCE) and KEYBIOEFFECTS (MRTN-CT-2006-035695). Additional funds were provided by
469 the SCARCE project (Consolider-Ingenio 2010, CSD2009-00065) of the Spanish Ministry of
470 Science and Innovation.

471 5. REFERENCES

472 Admiraal W, Blanck H, Buckert-De Jong M, Guasch H, Ivorra N, Lehmann V, Nyström, Paulsson M, Sabater S. Short-term toxicity of Zinc to microbenthic algae and bacteria in a metal polluted stream. Wat Res 1999; 3:1989-1996.

475 Adolfsson-Erici MM, Pettersson J, Parkkonen Sturve J. Triclosan, a commonly used bactericide 476 found in human milk and in the aquatic environment in Sweden. Chemosphere 2002; 46: 1485-1489.
477

478 Azevedo, D, Lacorte S, Vinhas T, Viana P, Barceló D. Monitoring of priority pesticides and other
479 organic pollutants in river water from Portugal by gas chromatography-mass
480 spectrometry and liquid chromatography-atmospheric pressure chemical ionization mass
481 spectrometry. J Chromatogr A 2000; 879: 13-26.
457 recovery capacity (by 2 weeks post-exposure they had recovered from nearly all the effects), but
458 also that even very short pulses of toxicants can have relevant consequences for biofilm structure
459 and function. It is reasonable to establish that short pulses can be seen as initial phases for effects
460 on biofilms, that longer toxicant pulses could imply more persistent effects and finally that
461 chronic concentrations of these toxicants could represent the most severe threat to biofilm
diversity and function.
462

463

464

465

466 Acknowledgements

467 This study was supported by the European Commission projects MODELKEY (Project 511237-2
468 GOCE) and KEYBIOEFFECTS (MRTN-CT-2006-035695). Additional funds were provided by
469 the SCARCE project (Consolider-Ingenio 2010, CSD2009-00065) of the Spanish Ministry of
470 Science and Innovation.

471 5. REFERENCES

472 Admiraal W, Blanck H, Buckert-De Jong M, Guasch H, Ivorra N, Lehmann V, Nyström, Paulsson M, Sabater S. Short-term toxicity of Zinc to microbenthic algae and bacteria in a metal polluted stream. Wat Res 1999; 3:1989-1996.

475 Adolfsson-Erici MM, Pettersson J, Parkkonen Sturve J. Triclosan, a commonly used bactericide 476 found in human milk and in the aquatic environment in Sweden. Chemosphere 2002; 46: 1485-1489.
477

478 Azevedo, D, Lacorte S, Vinhas T, Viana P, Barceló D. Monitoring of priority pesticides and other
479 organic pollutants in river water from Portugal by gas chromatography-mass
480 spectrometry and liquid chromatography-atmospheric pressure chemical ionization mass
481 spectrometry. J Chromatogr A 2000; 879: 13-26.
482 Battin TJ, Kaplan LA, Newbold JD, Hansen, CME. Contributions of microbial biofilms to
483 ecosystem processes in stream mesocosms. Nature 2003; 426 (6965): 439-442.

484 Blanchoud H, Farrugia F, Mouchel JM. Pesticide uses and transfers in urbanized catchments.
485 Chemosphere 2004; 55: 905-913.

486 Bonnineau C, Guasch H, Proia L, Ricart M, Geiszinger A, Romaní A.M, Sabater S.

487 Fluvial biofilms: A pertinent tool to assess β-blockers toxicity. Aquatic Toxicology 2010;
488 96: 225-233.

489 Bruckner CG, Kroth PG. Protocols for the removal of bacteria from freshwater benthic diatom
cultures. Journal of Phycology 2009; 45: 981-986.

491 Canesi L, Ciacci C, Lorusso LC, Betti M, Gallo G, Pojana G, Marcomini A. Effects of Triclosan
492 on Mytilus galloprovincialis hemocyte function and digestive gland enzyme activities:
493 Possible modes of action on non target organisms. Comparative Biochemistry and
494 Physiology 2007; 145(C): 464-472.

495 Capdevielle M, Van Egmond R, Whelan M, Versteeg D, Hofmann-Kamensky M, Inauen J,
496 Cunningham V, Woltering D. Consideration of Exposure and Species Sensitivity of
497 Triclosan in the Freshwater Environment. Integrated Environmental Assessment and
498 Management 2008; 4: 15-23.

499 Charles DF, Knowles C, Davies RS. Protocols for the analysis of algal samples collected as part
500 of the U.S. Geological Survey National Water-Quality Assessment Program. Patrick
501 Center for Environmental Research, The Academy of Natural Sciences; 2002,
502 Philadelphia, PA.

503 Chesworth JC, Donkin ME, Brown MT. The interactive effects of the antifouling herbicides
504 Irgarol 1051 and Diuron on the seagrass Zostera marina (L.). Aquatic Toxicology 2004;
505 66: 293-305.
482 Battin TJ, Kaplan LA, Newbold JD, Hansen, CME. Contributions of microbial biofilms to
483 ecosystem processes in stream mesocosms. Nature 2003; 426 (6965): 439-442.

484 Blanchoud H, Farrugia F, Mouchel JM. Pesticide uses and transfers in urbanized catchments.
485 Chemosphere 2004; 55: 905-913.

486 Bonnineau C, Guasch H, Proia L, Ricart M, Geiszinger A, Romaní A.M, Sabater S.

487 Fluvial biofilms: A pertinent tool to assess β-blockers toxicity. Aquatic Toxicology 2010;
488 96: 225-233.

489 Bruckner CG, Kroth PG. Protocols for the removal of bacteria from freshwater benthic diatom
cultures. Journal of Phycology 2009; 45: 981-986.

491 Canesi L, Ciacci C, Lorusso LC, Betti M, Gallo G, Pojana G, Marcomini A. Effects of Triclosan
492 on Mytilus galloprovincialis hemocyte function and digestive gland enzyme activities:
493 Possible modes of action on non target organisms. Comparative Biochemistry and
494 Physiology 2007; 145(C): 464-472.

495 Capdevielle M, Van Egmond R, Whelan M, Versteeg D, Hofmann-Kamensky M, Inauen J,
496 Cunningham V, Woltering D. Consideration of Exposure and Species Sensitivity of
497 Triclosan in the Freshwater Environment. Integrated Environmental Assessment and
498 Management 2008; 4: 15-23.

499 Charles DF, Knowles C, Davies RS. Protocols for the analysis of algal samples collected as part
500 of the U.S. Geological Survey National Water-Quality Assessment Program. Patrick
501 Center for Environmental Research, The Academy of Natural Sciences; 2002,
502 Philadelphia, PA.

503 Chesworth JC, Donkin ME, Brown MT. The interactive effects of the antifouling herbicides
504 Irgarol 1051 and Diuron on the seagrass Zostera marina (L.). Aquatic Toxicology 2004;
505 66: 293-305.
506 Croft MT, Lawrence AD, Raux-Deery E, Warren MJ, Smith AG. Algae acquire vitamin B12
507 through a symbiotic relationship with bacteria. Nature 2005; 438:90-3.

508 De Lorenzo ME, Keller JM, Arthur CD, Finnegan MC, Harper HE, Winder VL, Zdankiewicz
509 DL. Toxicity of the Antimicrobial Compound Triclosan and Formation of the Metabolite
510 Methyl-Triclosan in Estuarine Systems. Environmental Toxicology 2007; 23: 224-232.

511 Ellis JB. Pharmaceutical and personal care products (PPCPs) in urban receiving waters.
512 Environmental Pollution 2007; 144: 184-189.

513 European Commission. Directive 2000/60/EC of the European Parliament and of the Council –
514 Establishing a Framework for Community. Action in the Field of Water Policy 2000;
515 European Commission, Belgium.

516 Escalada M, Russell AD, Maillard JY, Ochs D. Triclosan-bacteria interactions: single or multiple
517 target sites? Letters in Applied Microbiology 2005; 41: 476-481.

518 Farré M, Asperger D, Kantiani L, González S, Petrovic M, Barceló D. Assessment of the acute
519 toxicity of triclosan and methyl triclosan in wastewater based on the bioluminescence
520 inhibition of Vibrio fischeri. Anal Bioanal Chem. 2008; 390:1999-2007.

521 Findlay S, Strayer D, Goumbala C, Gould K. Metabolism of streamwater dissolved organic
522 carbon in the shallow hyporheic zone. Limnology and Oceanography 1993; 38: 1493-1499.

523 Flaherty CM, Dodson SI. Effects of pharmaceuticals on Daphnia survival, growth, and
524 reproduction. Chemosphere 2005; 61: 200-207.

525 Francoeur SN, Wetzel RG. Regulation of periphytic leucine-aminopeptidase activity. Aquatic
526 Microbial Ecology 2003; 31: 249-258.

527 Franz S, Altenburger R, Heilmeier H, Schmitt-Janse M. What contributes to the sensitivity of
528 microalgae to triclosan? Aquatic Toxicology 2008; 90: 102-108.

529 Giacomazzi S, Cochet N. Environmental impact of diuron transformation: a review.
530 Chemosphere 2004; 56: 1021-1032.
506 Croft MT, Lawrence AD, Raux-Deery E, Warren MJ, Smith AG. Algae acquire vitamin B12
507 through a symbiotic relationship with bacteria. Nature 2005; 438:90-3.

508 De Lorenzo ME, Keller JM, Arthur CD, Finnegan MC, Harper HE, Winder VL, Zdankiewicz
509 DL. Toxicity of the Antimicrobial Compound Triclosan and Formation of the Metabolite
510 Methyl-Triclosan in Estuarine Systems. Environmental Toxicology 2007; 23: 224-232.

511 Ellis JB. Pharmaceutical and personal care products (PPCPs) in urban receiving waters.
512 Environmental Pollution 2007; 144: 184-189.

513 European Commission. Directive 2000/60/EC of the European Parliament and of the Council -
514 Establishing a Framework for Community. Action in the Field of Water Policy 2000;
515 European Commission, Belgium.

516 Escalada M, Russell AD, Maillard JY, Ochs D. Triclosan-bacteria interactions: single or multiple
517 target sites? Letters in Applied Microbiology 2005; 41: 476-481.

518 Farré M, Asperger D, Kantiani L, González S, Petrovic M, Barceló D. Assessment of the acute
519 toxicity of triclosan and methyl triclosan in wastewater based on the bioluminescence
520 inhibition of Vibrio fischeri. Anal Bioanal Chem. 2008; 390:1999-2007.

521 Findlay S, Strayer D, Goumbala C, Gould K. Metabolism of streamwater dissolved organic
522 carbon in the shallow hyporheic zone. Limnology and Oceanography 1993; 38: 1493-1499.

523 Flaherty CM, Dodson SI. Effects of pharmaceuticals on Daphnia survival, growth, and
524 reproduction. Chemosphere 2005; 61: 200-207.

525 Francoeur SN, Wetzel RG. Regulation of periphytic leucine-aminopeptidase activity. Aquatic
526 Microbial Ecology 2003; 31: 249-258.

527 Franz S, Altenburger R, Heilmeier H, Schmitt-Janse M. What contributes to the sensitivity of
528 microalgae to triclosan? Aquatic Toxicology 2008; 90: 102-108.

529 Giacomazzi S, Cochet N. Environmental impact of diuron transformation: a review.
530 Chemosphere 2004; 56: 1021-1032.
531 Guillard RRL. Division rates. In: Stein JR editor. Culture methods and growth measurements.

532 Handbook of phycological methods. Cambridge University Press, Cambridge; 1973. p.

533 289-311.

534 Halden R, Paull DH. Co-occurrence of triclocarban and triclosan in U.S. water resources. Environ
535 Sci Technol. 2005; 39:1420-1426.

536 Kantiani L, Farré M, Asperger D, Rubio F, González S, López de Alda M, Petrović M, Shelver
537 W, Barceló D. Triclosan ans methyl-triclosan study in the northeast of Spain using
538 magnetic particle enzyme immunoassay and confirmatory analysis by gas cromatography-
539 mass spectrometry. Journal of Hydrology 2008; 361:1-9.

540 Krammer K. Lange-Bertalot H. Bacillariophyceae 1. Teil: Naviculaceae. p. 876; 2. Teil:
541 Bacillariaceae, Epithemiaceae, Surirellaceae, p. 596; 3. Teil: Centrales, Fragilariaceae,
542 Eunotiaceae, p. 576; 4. Teil: Achnanthaceae. Kritische Ergänzungen zu Navicula
543 (Lineolatae) und Gomphonema. p. 437 G. Fischer Verlag., Stuttgart. 1986 – 1991.

544 Kuster M, López de Alda M, Hernando MD, Petrovic M, Martín-Alonso J, Barceló D. Analysis
545 and occurrence of pharmaceuticals, estrogens, progestogens and polar pesticides in
546 sewage treatment plant effluents, river water and drinking water in the Llobregat river
547 basin (Barcelona, Spain). Journal of Hydrology 2008; 358: 1123-123.

548 Jeffrey S, Humphrey GF. New spectrophotometric equations for determining chlorophylls a, b,
549 c1 and c2 in higher plants, algae and natural phytoplankton. Biochemie und Physiologie
der Pflanzen 1975; 167: 191-194.

551 Lawrence JR, Zhu B, Swerhone GDW, Roy J, Wassenaar LI, Topp E, Korber DR. Comparative
552 microscale analysis of the effects of triclosan and triclocarban on the structure and
553 function of river biofilm communities. Science of the Total Environment 2009; 407:
554 3307-3316.
531 Guillard RRL. Division rates. In: Stein JR editor. Culture methods and growth measurements.

532 Handbook of phycological methods. Cambridge University Press, Cambridge; 1973. p.

533 289-311.

534 Halden R, Paull DH. Co-occurrence of triclocarban and triclosan in U.S. water resources. Environ
535 Sci Technol. 2005; 39:1420-1426.

536 Kantiani L, Farré M, Asperger D, Rubio F, González S, López de Alda M, Petrović M, Shelver
537 W, Barceló D. Triclosan ans methyl-triclosan study in the northeast of Spain using
538 magnetic particle enzyme immunoassay and confirmatory analysis by gas cromatography-
539 mass spectrometry. Journal of Hydrology 2008; 361:1-9.

540 Krammer K. Lange-Bertalot H. Bacillariophyceae 1. Teil: Naviculaceae. p. 876; 2. Teil:
541 Bacillariaceae, Epithemiaceae, Surirellaceae, p. 596; 3. Teil: Centrales, Fragilariaceae,
542 Eunotiaceae, p. 576; 4. Teil: Achnanthaceae. Kritische Ergänzungen zu Navicula
543 (Lineolatae) und Gomphonema. p. 437 G. Fischer Verlag., Stuttgart. 1986 – 1991.

544 Kuster M, López de Alda M, Hernando MD, Petrovic M, Martín-Alonso J, Barceló D. Analysis
545 and occurrence of pharmaceuticals, estrogens, progestogens and polar pesticides in
546 sewage treatment plant effluents, river water and drinking water in the Llobregat river
547 basin (Barcelona, Spain). Journal of Hydrology 2008; 358: 1123-123.

548 Jeffrey S, Humphrey GF. New spectrophotometric equations for determining chlorophylls a, b,
549 c1 and c2 in higher plants, algae and natural phytoplankton. Biochemie und Physiologie
der Pflanzen 1975; 167: 191-194.

551 Lawrence JR, Zhu B, Swerhone GDW, Roy J, Wassenaar LI, Topp E, Korber DR. Comparative
552 microscale analysis of the effects of triclosan and triclocarban on the structure and
553 function of river biofilm communities. Science of the Total Environment 2009; 407:
554 3307-3316.
555 Legrand H, Herlory O, Guarini J, Blanchard GF, Richard P. Inhibition of microphytobenthic
556 photosynthesis by the herbicides atrazine and diuron. Cahiers de Biologie Marine 2006;
557 47: 39-45.

558 Lock MA. Attached microbial communities in rivers. In: Fored TE (editor) Aquatic
559 Microbiology: an Ecological Approach. Blackwell Scientific Publications, Oxford 1993;
560 p. 113–138.

561 Loos R, Wollgast J, Huber T, Hanke G. Polar herbicides, pharmaceutical products,
562 perfluorooctanesulfonate (PFOS), perfluorooctanoate (PFOA), and nonylphenol and its
563 carboxylates and ethoxylates in surface and tap waters around Lake Maggiore in Northern
564 Italy. Anal Bioanal Chem. 2007; 387: 1469–1478.

565 López-Doval JC, Ricart M, Guasch H, Romaní AM, Sabater S, Muñoz I. Does Grazing Pressure
566 Modify Diuron Toxicity in a Biofilm Community? Archives of Environmental
567 Contamination and Toxicology 2010; 58: 955-962.

568 Mathuriau C, Chauvet E. Breakdown of litter in a neotropical stream. J. N. Am. Benthol. Soc.
569 2002; 21: 384-396.

570 McAvoy, DC, Schatowitz B, Jacob M, Hauk A, Eckhoff WS. Measurement of triclosan in
571 wastewater treatment systems. Environmental Toxicology and Chemistry 2002; 21: 1323–
572 1329.

573 McMurry LM, Oethinger M, Levy SB. Triclosan targets lipid synthesis. Nature 1998; 394: 531-
574 532.

575 Mølander S, Blanck H. Detection of pollution-induced community tolerance
576 (PICT) in marine periphyton communities established under diuron exposure. Aquatic
577 Toxicology 1992; 22: 129–144.

578 Morin S, Coste M, Delmas F. A comparison of specific growth rates of periphytic diatoms of
579 varying cell size under laboratory and field conditions. Hydrobiologia 2008; 614: 285–
580 297.
555 Legrand H, Herlory O, Guarini J, Blanchard GF, Richard P. Inhibition of microphytobenthic photosynthesis by the herbicides atrazine and diuron. Cahiers de Biologie Marine 2006; 47: 39-45.

558 Lock MA. Attached microbial communities in rivers. In: Fored TE (editor) Aquatic Microbiology: an Ecological Approach. Blackwell Scientific Publications, Oxford 1993; p. 113-138.

561 Loos R, Wollgast J, Huber T, Hanke G. Polar herbicides, pharmaceutical products, perfluorooctanesulfonate (PFOS), perfluorooctanoate (PFOA), and nonylphenol and its carboxylates and ethoxylates in surface and tap waters around Lake Maggiore in Northern Italy. Anal Bioanal Chem. 2007; 387: 1469-1478.

565 López-Doval JC, Ricart M, Guasch H, Romaní AM, Sabater S, Muñoz I. Does Grazing Pressure Modify Diuron Toxicity in a Biofilm Community? Archives of Environmental Contamination and Toxicology 2010; 58: 955-962.

568 Mathuriau C, Chauvet E. Breakdown of litter in a neotropical stream. J. N. Am. Benthol. Soc. 2002; 21: 384-396.

570 McAvoy, DC, Schatowitz B, Jacob M, Hauk A, Eckhoff WS. Measurement of triclosan in wastewater treatment systems. Environmental Toxicology and Chemistry 2002; 21: 1323-1329.

573 McMurry LM, Oethinger M, Levy SB. Triclosan targets lipid synthesis. Nature 1998; 394: 531-532.

575 Mølander S, Blanck H. Detection of pollution-induced community tolerance (PICT) in marine periphyton communities established under diuron exposure. Aquatic Toxicology 1992; 22: 129-144.

578 Morin S, Coste M, Delmas F. A comparison of specific growth rates of periphytic diatoms of varying cell size under laboratory and field conditions. Hydrobiologia 2008; 614: 285-297.
581 Morin S, Pesce S, Tlili A, Coste M, Montuelle B. Recovery potential of periphytic communities
582 in a river impacted by a vineyard watershed. Ecological Indicators 2010a; 10: 419-426.

583 Morin S, Proia L, Ricart M, Bonnineau C, Geiszinger A, Ricciardi F, Guasch H, Romani AM,
584 Sabater S. Effects of a bactericide on the structure and survival of benthic diatom
585 communities. Life and Environment 2010b; 60: 107-114.

586 Morrall D, McAvoy D, Schatowitz B, Inauen J, Jacob M, Hauk A, Eckhoff W. A field study of
587 triclosan loss rates in river water (Cibolo Creek, TX). Chemosphere 2004; 54: 653-660.

588 Orvos DR, Vergsteeg DJ, Inauen J, Capdeville M, Rothrnstein A, Cunnigham V. Aquatic toxicity
589 of Triclosan. Environmental Toxicology and Chemistry 2002; 21: 1338-1349.

590 Pesce S, Fajon C, Bardot C, Bonnemoy F, Portelli C, Bohatier J. Effects of the phenylurea
591 herbicide diuron on natural riverine microbial communities in an experimental study.
592 Aquatic Toxicology 2006; 78: 303-314.

593 Pesce S, Bardot C, Lehours A, Batisson I, Bohatier J, Fajon C. Effects of diuron in microcosms
594 on natural riverine bacterial community composition: new insight into phylogenetic
595 approaches using PCR-TTGE analysis. Aquatic Sciences 2008; 70: 410-418.

596 Phan T, Marquis RE. Triclosan inhibition of membrane enzymes and glycolysis of *Streptococcus mutans* in suspension and biofilm. Canadian Journal of Microbiology 2006; 52: 977-983.

598 Rabiet M, Margoum C, Gouy V, Carluer N, Coquery M. Assessing pesticide concentrations and
599 fluxes in the stream of a small vineyard catchment - Effect of sampling frequency.
600 Environmental Pollution 2010; 158: 737-748.

601 Ricart M, Barceló D, Geiszinger A, Guasch H, López de Alda M, Romaní AM, Vidal G,
602 Villagrasa M, Sabater S. Effects of low concentrations of the phenylurea herbicide diuron
603 on biofilm algae and bacteria. Chemosphere 2009; 76: 1392-1401.

604 Ricart M, Guasch H, Alberch M, Barceló D, Bonnineau C, Geiszinger A, Farré M, Ferrer J,
605 Ricciardi F, Romaní A, Morin S, Proia L, Sala L, Sureda D, Sabater S. Triclosan
581 Morin S, Pesce S, Tlili A, Coste M, Montuelle B. Recovery potential of periphytic communities
582 in a river impacted by a vineyard watershed. Ecological Indicators 2010a; 10: 419-426.

583 Morin S, Proia L, Ricart M, Bonnineau C, Geiszinger A, Ricciardi F, Guasch H, Romani AM,
584 Sabater S. Effects of a bactericide on the structure and survival of benthic diatom
585 communities. Life and Environment 2010b; 60: 107-114.

586 Morrall D, McAvoy D, Schatowitz B, Inauen J, Jacob M, Hauk A, Eckhoff W. A field study of
587 triclosan loss rates in river water (Cibolo Creek, TX). Chemosphere 2004; 54: 653-660.

588 Orvos DR, Vergsteeg DJ, Inauen J, Capdeville M, Rothrnstein A, Cunnigham V. Aquatic toxicity
589 of Triclosan. Environmental Toxicology and Chemistry 2002; 21: 1338-1349.

590 Pesce S, Fajon C, Bardot C, Bonnemoy F, Portelli C, Bohatier J. Effects of the phenylurea
591 herbicide diuron on natural riverine microbial communities in an experimental study.
592 Aquatic Toxicology 2006; 78: 303-314.

593 Pesce S, Bardot C, Lehours A, Batisson I, Bohatier J, Fajon C. Effects of diuron in microcosms
594 on natural riverine bacterial community composition: new insight into phylogenetic
595 approaches using PCR-TTGE analysis. Aquatic Sciences 2008; 70: 410-418.

596 Phan T, Marquis RE. Triclosan inhibition of membrane enzymes and glycolysis of *Streptococcus mutans* in suspension and biofilm. Canadian Journal of Microbiology 2006; 52: 977-983.

598 Rabiet M, Margoum C, Gouy V, Carluer N, Coquery M. Assessing pesticide concentrations and
599 fluxes in the stream of a small vineyard catchment - Effect of sampling frequency.
600 Environmental Pollution 2010; 158: 737-748.

601 Ricart M, Barceló D, Geiszinger A, Guasch H, López de Alda M, Romaní AM, Vidal G,
602 Villagrasa M, Sabater S. Effects of low concentrations of the phenylurea herbicide diuron
603 on biofilm algae and bacteria. Chemosphere 2009; 76: 1392-1401.

604 Ricart M, Guasch H, Alberch M, Barceló D, Bonnineau C, Geiszinger A, Farré M, Ferrer J,
605 Ricciardi F, Romaní A, Morin S, Proia L, Sala L, Sureda D, Sabater S. Triclosan
606 persistence trough wastewater treatment plants and its potential toxic effects on river
607 biofilms. Aquatic Toxicology 2010; 100: 346-353.

608 Rodriguez-Mozaz S, López de Alda MJ, Barceló D. Monitoring of estrogens, pesticides and
609 bisphenol A in natural waters and drinking water treatment plants by solid-phase
610 extraction-liquid chromatography-mass spectrometry. J. Chromatogr. 2004; 1045(A):
611 85-92.

612 Romaní AM, Sabater S. Effect of primary producers on the heterotrophic metabolism of a stream
613 biofilm. Freshwater Biology 1999; 41: 729-736.

614 Sabater S, Tockner K. Effects of hydrologic alterations on the ecological quality of river
615 ecosystems. In: Sabater S and Barceló D (editors). Water scarcity in the Mediterranean:
616 Perspectives Under Global Change. Springer Verlag. Hdb Env. Chem. 2010; 8: 15-39.

617 Sabater S, Guasch H, Ricart M, Romaní AM, Vidal G, Klünder C, Schmitt- Jansen M.
618 Monitoring the effect of chemicals on biological communities. The biofilm as an
619 interface. Anal. Bioanal. Chem. 2007; 387, 1425-1434.

620 Samrakandi MM, Roques C, Michel G. Influence of trophic conditions on exopolysaccharide
621 production: bacterial biofilm susceptibility to chlorine and monochloramine. Can J
622 Microbiol 1997; 3: 751-758.

623 Samsøe-Petersen L, Winther-Nielsen M, Madsen T. Fate and Effects of Triclosan. Environmental
624 Project No. 861 2003. Miljøprojekt. Danish Environmental Protection Agency. Danish
625 Ministry of The Environment, 2003.

626 Serra A, Corcoll N, Guasch H. Copper accumulation and toxicity in fluvial periphyton: the
627 influence of exposure history. Chemosphere 2009; 74: 633-641.

628 Singer H, Muller S, Tixier C, Pillonel L. Triclosan: occurrence and fate of a widely used biocide
629 in the aquatic environment: field measurements in wastewater treatment plants, surface
630 waters, and lake sediments. Environ. Sci. Technol. 2002; 36: 4998-5004.
606 persistence trough wastewater treatment plants and its potential toxic effects on river
607 biofilms. Aquatic Toxicology 2010; 100: 346-353.

608 Rodriguez-Mozaz S, López de Alda MJ, Barceló D. Monitoring of estrogens, pesticides and
609 bisphenol A in natural waters and drinking water treatment plants by solid-phase
610 extraction-liquid chromatography-mass spectrometry. J. Chromatogr. 2004; 1045(A):
611 85-92.

612 Romaní AM, Sabater S. Effect of primary producers on the heterotrophic metabolism of a stream
613 biofilm. Freshwater Biology 1999; 41: 729-736.

614 Sabater S, Tockner K. Effects of hydrologic alterations on the ecological quality of river
615 ecosystems. In: Sabater S and Barceló D (editors). Water scarcity in the Mediterranean:
616 Perspectives Under Global Change. Springer Verlag. Hdb Env. Chem. 2010; 8: 15-39.

617 Sabater S, Guasch H, Ricart M, Romaní AM, Vidal G, Klünder C, Schmitt- Jansen M.
618 Monitoring the effect of chemicals on biological communities. The biofilm as an
619 interface. Anal. Bioanal. Chem. 2007; 387, 1425-1434.

620 Samrakandi MM, Roques C, Michel G. Influence of trophic conditions on exopolysaccharide
621 production: bacterial biofilm susceptibility to chlorine and monochloramine. Can J
622 Microbiol 1997; 3: 751-758.

623 Samsøe-Petersen L, Winther-Nielsen M, Madsen T. Fate and Effects of Triclosan. Environmental
624 Project No. 861 2003. Miljøprojekt. Danish Environmental Protection Agency. Danish
625 Ministry of The Environment, 2003.

626 Serra A, Corcoll N, Guasch H. Copper accumulation and toxicity in fluvial periphyton: the
627 influence of exposure history. Chemosphere 2009; 74: 633-641.

628 Singer H, Muller S, Tixier C, Pillonel L. Triclosan: occurrence and fate of a widely used biocide
629 in the aquatic environment: field measurements in wastewater treatment plants, surface
630 waters, and lake sediments. Environ. Sci. Technol. 2002; 36: 4998-5004.
631 Tatarazako N, Hishibashi H, Teshima K, Kishi K, Arizono K. Effects of Triclosan on Various

632 Aquatic Organisms. Environmental Sciences 2004; 11: 133-140.

633 Tlili A, Dorigo U, Montuelle B, Margoum C, Carluer N, Gouy V, Bouchez A, Bérard A.

634 Responses of chronically contaminated biofilms to short pulses of diuron. An

635 experimental study simulating flooding events in a small river. Aquatic Toxicology 2008;

636 87: 252-263.

637 Van Rensen JJS. Herbicides interacting with photosystem II. In: Dodge, A.D. (Ed.), Herbicides

638 and Plant Metabolism. Cambridge University Press, Cambridge 1989; 21-36.

639 Villalaín J, Reyes Mateo C, Aranda FJ, Shapiro S, Micol V. Membranotropic Effects of the

640 Antibacterial Agent Triclosan. Archives of Biochemistry and Biophysics 2001; 390: 128-

641 136.

642 White KL, Haggard BE, Matlock MD, Kim MD. Periphytic chlorophyll-a response to triclosan

643 exposure: application of a passive diffusion periphytometer. Applied Engineering in

644 Agriculture 2005; 21: 307-311.

645 Wilson BA, Smith V, Denoyelles F Jr, Larive CK. Effects of Three Pharmaceutical and Personal

646 Care Products on Natural Freshwater Algal Assemblages. Environmental Science &

647 Technology 2003; 37: 1713-1719.

648 Xie Z, Ebinghaus R, Flöser G, Caba A, Ruck W. Occurrence and distribution of triclosan in the

649 German Bight (North Sea). Environmental Pollution 2008; 156: 1190-1195.

650 Ylla I, Borrego C, Romaní AM, Sabater S. Availability of glucose and light modulates the

651 structure and function of a microbial biofilm. FEMS Microbiology Ecology 2009; 69: 27-

652 42.

653 FIGURE CAPTIONS

654 1) Changes of control (black) and treated (greys) biofilms in following endpoints: (a) Chla

655 density; (b) live/dead diatom ratio and (c) live/dead bacteria ratio in each sampling date.

656 . Values are means and standard deviation (n = 4). Post-hoc Tukey-b analysis results are
631 Tatarazako N, Hishibashi H, Teshima K, Kishi K, Arizono K. Effects of Triclosan on Various

632 Aquatic Organisms. Environmental Sciences 2004; 11: 133-140.

633 Tlili A, Dorigo U, Montuelle B, Margoum C, Carluer N, Gouy V, Bouchez A, Bérard A.

634 Responses of chronically contaminated biofilms to short pulses of diuron. An

635 experimental study simulating flooding events in a small river. Aquatic Toxicology 2008;

636 87: 252-263.

637 Van Rensen JJS. Herbicides interacting with photosystem II. In: Dodge, A.D. (Ed.), Herbicides

638 and Plant Metabolism. Cambridge University Press, Cambridge 1989; 21-36.

639 Villalaín J, Reyes Mateo C, Aranda FJ, Shapiro S, Micol V. Membranotropic Effects of the

640 Antibacterial Agent Triclosan. Archives of Biochemistry and Biophysics 2001; 390: 128-

641 136.

642 White KL, Haggard BE, Matlock MD, Kim MD. Periphytic chlorophyll-a response to triclosan

643 exposure: application of a passive diffusion periphytometer. Applied Engineering in

644 Agriculture 2005; 21: 307-311.

645 Wilson BA, Smith V, Denoyelles F Jr, Larive CK. Effects of Three Pharmaceutical and Personal

646 Care Products on Natural Freshwater Algal Assemblages. Environmental Science &

647 Technology 2003; 37: 1713-1719.

648 Xie Z, Ebinghaus R, Flöser G, Caba A, Ruck W. Occurrence and distribution of triclosan in the

649 German Bight (North Sea). Environmental Pollution 2008; 156: 1190-1195.

650 Ylla I, Borrego C, Romaní AM, Sabater S. Availability of glucose and light modulates the

651 structure and function of a microbial biofilm. FEMS Microbiology Ecology 2009; 69: 27-

652 42.

653 FIGURE CAPTIONS

654 1) Changes of control (black) and treated (greys) biofilms in following endpoints: (a) Chla

655 density; (b) live/dead diatom ratio and (c) live/dead bacteria ratio in each sampling date.

656 . Values are means and standard deviation (n = 4). Post-hoc Tukey-b analysis results are
657 showed when treatment effect resulted significant. Statistical significance was set at $p \le 0.05$ (one-way ANOVA).

659 2) Diatom community evolution during experiment. a) Increase of live cell density from day 2 to day 16. Slope of linear curve is the grow rate (div day⁻¹) of diatoms

661 3) SEM pictures of Control (1), DIU-treated (2) and TCS-treated (3) biofilms at day 16.

662 4) Changes of control (black) and treated (greys) biofilms in community photosynthetic efficiency; in each sampling date. Values are means and standard deviation (n = 4). Post-hoc Tukey-b analysis results are showed when treatment effect resulted significant. Statistical significance was set at $p \le 0.05$ (one-way ANOVA).

666 5) Changes in Phosphate Uptake rates (U) of control (black) and treated (greys) biofilms in each sampling date. Values are means and standard deviation (n = 4). Post-hoc Tukey-b analysis results are showed when treatment effect resulted significant. Statistical significance was set at $p \le 0.05$ (one-way ANOVA).

670
657 showed when treatment effect resulted significant. Statistical significance was set at $p \le 0.05$ (one-way ANOVA).

659 2) Diatom community evolution during experiment. a) Increase of live cell density from day 2 to day 16. Slope of linear curve is the grow rate (div day⁻¹) of diatoms

661 3) SEM pictures of Control (1), DIU-treated (2) and TCS-treated (3) biofilms at day 16.

662 4) Changes of control (black) and treated (greys) biofilms in community photosynthetic efficiency; in each sampling date. Values are means and standard deviation (n = 4). Post-hoc Tukey-b analysis results are showed when treatment effect resulted significant. Statistical significance was set at $p \le 0.05$ (one-way ANOVA).

666 5) Changes in Phosphate Uptake rates (U) of control (black) and treated (greys) biofilms in each sampling date. Values are means and standard deviation (n = 4). Post-hoc Tukey-b analysis results are showed when treatment effect resulted significant. Statistical significance was set at $p \le 0.05$ (one-way ANOVA).

670