**Chlorophyll-a and the Supply Side Ecology: Lessons from the Rocky Shores**

Ana Carolina de Azevedo Mazzuco and Paula Kasten

Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/68044

#### **Abstract**

The aims of this study were to summarize and describe the influences of phytoplankton on the larval cycle of rocky shore invertebrates, and to assess the relationship between fluctuations in chlorophyll-a concentration and the rates of larval processes. We carried out a mini review of the published data regarding the theme of the chapter, in which we described the ecological trends for the most common taxa and key species at small and larger spatiotemporal scales. The following topics were addressed: (i) the influence of phytoplankton on larval development, rhythms of larval release, larval quality, larval transport, settlement, and recruitment; (ii) the relationships between variations in chlorophyll-a concentration and the rates of larval processes; (iii) climate change on phytoplankton larva dynamics. The information presented here highlights the role of phytoplankton on rocky shore communities, as well as the importance of chlorophyll-a as a tool for modeling and forecasting the supply side ecology in rocky shore communities.

**Keywords:** phytoplankton, chlorophyll-a, supply side ecology, marine invertebrates, rocky shores, benthic-pelagic coupling

#### **1. Introduction**

Larval supply is the main source of new individuals to the populations of rocky shore invertebrates [1–3]. In these communities, larval success regulates how energy is transferred through the trophic web [4–6]; consequently, variations in the supply of propagules are the basis of trophic interactions at rocky shores [7, 8]. Since phytoplankton is the main food source for planktonic larvae of marine invertebrates [9], variations in phytoplankton biomass and diversity have significant influences on the larval cycle. Larval responses to the variability in phytoplankton abundance and diversity are species-specific. Larval fitness is influenced by environmental conditions experienced by adults

© 2017 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

and larvae [10, 11]. The effects of phytoplankton on larval dynamics depend on the phase of larval development [12–15] and may be stronger when variations in phytoplankton occur on temporal scales that larvae or breeding adults are able to respond [16]. The direct interaction between phytoplankton and the larval stages have short-term consequences for larval dynamics (e.g., Ref. [14]), and it might have long-term effects as well. Because of that, variations in the rates of the ecological processes of rocky shore invertebrates are commonly correlated with fluctuations in chlorophyll-a concentration in the ocean (e.g., Refs. [17–20]). These numerical relationships are important tools to ecological modeling, and may be used to improve stock management in some extent [21].

#### **2. The role of phytoplankton blooms in reproduction timing and in the rhythms of larval release**

In the rocky shore communities, filter feeders depend greatly on phytoplankton as their main source of food and its consumption results in energy for growth and reproduction [22]. It is common to find larger animals with higher fecundity rates at rocky shores located in areas of high primary productivity, as a response to the higher concentrations of phytoplankton, and thus, food availability [19, 23–25]. Different types of phytoplankton present distinct physiological qualities as food particles [26], thereby both the amount of phytoplankton in the water column and their diversity influence the reproductive traits in marine invertebrates.

But not only adults on the rocky shore depend on phytoplankton in order to survive, larvae produced by those organisms also rely on these microorganisms to develop and reach the juvenile phase [27]. As evolution drives maximum reproductive activity to happen when environmental conditions are the best for offspring development, food availability is one of the most important factors regulating reproduction and allowing adults to produce viable offspring. Thus, it is common to observe peaks of larval release by rocky shore invertebrates synchronized with phytoplankton blooms (e.g., Refs. [28, 29]). Some metabolites produced by phytoplankton are signs of favorable environmental conditions for the larval development, trigging the spawning activity of green sea urchins and blue mussels, for instance Ref. [28]. These animals perceive such chemical compounds as an indication of good food abundance, so synchronizing the timing of larval release with high abundance of phytoplankton would promote higher offspring survival. Barnacles, on the other hand, just need a physical contact with phytoplankton cells to trigger their spawning activity, and larger the phytoplankton cell is, the stronger is the response [28].

Therefore, the presence of phytoplankton may overcome other environmental factors in the regulation of reproduction timing and larval release [30]. Spring and summer are the main reproductive periods for rocky shore invertebrates at temperate and upwelling regions [31], as it is during these seasons that phytoplankton blooms occur. Mussels from the Baltic sea, for example, start to develop their gonads when temperature starts to drop in the beginning of winter; but its maturation and ripening processes proceed in a way that the animals are ready to reproduce at the same time that phytoplankton blooms occur in the beginning of spring [32]. Some barnacles are even able to maintain their fully developed nauplii in the mantle cavity until a high abundance of phytoplankton is perceived by the adults and only then, the nauplii will be released, a strategy that enhances the offspring survival due to the higher chance of facing a favorable feeding environment [33].

and larvae [10, 11]. The effects of phytoplankton on larval dynamics depend on the phase of larval development [12–15] and may be stronger when variations in phytoplankton occur on temporal scales that larvae or breeding adults are able to respond [16]. The direct interaction between phytoplankton and the larval stages have short-term consequences for larval dynamics (e.g., Ref. [14]), and it might have long-term effects as well. Because of that, variations in the rates of the ecological processes of rocky shore invertebrates are commonly correlated with fluctuations in chlorophyll-a concentration in the ocean (e.g., Refs. [17–20]). These numerical relationships are important tools to ecological modeling, and may be used to improve stock management in some extent [21].

**2. The role of phytoplankton blooms in reproduction timing and in the** 

In the rocky shore communities, filter feeders depend greatly on phytoplankton as their main source of food and its consumption results in energy for growth and reproduction [22]. It is common to find larger animals with higher fecundity rates at rocky shores located in areas of high primary productivity, as a response to the higher concentrations of phytoplankton, and thus, food availability [19, 23–25]. Different types of phytoplankton present distinct physiological qualities as food particles [26], thereby both the amount of phytoplankton in the water column and their diversity influence the reproductive traits in marine invertebrates.

But not only adults on the rocky shore depend on phytoplankton in order to survive, larvae produced by those organisms also rely on these microorganisms to develop and reach the juvenile phase [27]. As evolution drives maximum reproductive activity to happen when environmental conditions are the best for offspring development, food availability is one of the most important factors regulating reproduction and allowing adults to produce viable offspring. Thus, it is common to observe peaks of larval release by rocky shore invertebrates synchronized with phytoplankton blooms (e.g., Refs. [28, 29]). Some metabolites produced by phytoplankton are signs of favorable environmental conditions for the larval development, trigging the spawning activity of green sea urchins and blue mussels, for instance Ref. [28]. These animals perceive such chemical compounds as an indication of good food abundance, so synchronizing the timing of larval release with high abundance of phytoplankton would promote higher offspring survival. Barnacles, on the other hand, just need a physical contact with phytoplankton cells to trigger their spawning activity, and larger the phytoplankton cell

Therefore, the presence of phytoplankton may overcome other environmental factors in the regulation of reproduction timing and larval release [30]. Spring and summer are the main reproductive periods for rocky shore invertebrates at temperate and upwelling regions [31], as it is during these seasons that phytoplankton blooms occur. Mussels from the Baltic sea, for example, start to develop their gonads when temperature starts to drop in the beginning of winter; but its maturation and ripening processes proceed in a way that the animals are ready to reproduce at the same time that phytoplankton blooms occur in the beginning of spring [32]. Some barnacles are even able to maintain their fully developed nauplii in the mantle

**rhythms of larval release**

8 Chlorophyll

is, the stronger is the response [28].

Similar reproductive timing was registered in the Indian coasts, where phytoplankton blooms occur during the monsoons and barnacles spawn their nauplii short after a break of the monsoon conditions [34]. However, these are not the best conditions for nauplii development, as these breaks stop and unfavorable monsoon conditions for larval development return soon after. Such misleading cue could result in lower recruitment rates for barnacles in this region. In subtropical coasts, peaks of larval production in intertidal barnacles are also preceded by high concentrations of chlorophyll-a in the water column [35]. On the daily scale, phytoplankton diversity might be as important as biomass in the regulation of larval release [36]. The presence of phytoplankton may overcome other environmental factors known to act as synchronization cues for reproduction timing and larval release [30].

#### **3. How do changes in phytoplankton affect larval development from release to competency?**

As seen in the previous section, phytoplankton has an important role in the reproductive success of marine invertebrates inhabiting the rocky shores. Part of this reproductive success involves the survival of larvae up to the juvenile stage, and a successful return to the benthic habitat is essential to the maintenance of rocky shore populations [2, 37]. It is straightforward to think that larval development is strictly linked to changes in phytoplankton community, since these cells are the main food items for marine planktotrophic larvae [9]. Because of that, the physiological quality of a larva would be determined in the plankton during its development and influenced directly by the phytoplankton in the water column. However, phytoplankton may change larval physiological quality much before that same larva is produced, through maternal effects, that is, when maternal individuals have the capacity to perceive the environment and manipulate the energy allocated for propagule production [38].

The amount of energetic reserves allocated to each propagule produced depends on the amount of energy the maternal individual can provide to its offspring. This capacity, in turn, is limited by the food available for the mothers, their perception of it, and their competency to gather and assimilate energy [38, 39]. For those marine organisms that produce lecithotrophic larvae, maternal effects are extremely important for shaping larval physiological quality because these larvae depend exclusively on the energetic resources from embryogenesis to survive [40]. If food ration is low, mothers can either preserve the energy acquired for their own metabolism and produce lower quality larvae (a selfish strategy, Ref. [39]) or invest all energy possible into their propagules, enhancing the survival potential of that higher quality larvae (an anticipatory strategy, Ref. [39]). In a scenario where maternal individuals are feeding mainly on phytoplankton, as the majority of filter feeding invertebrates in the rocky shores are, it is possible to understand the effect that oscillations in the quantity and type of phytoplankton available for these animals to feed has on larval quality.

However, most invertebrates that inhabit the rocky shores produce planktotrophic larvae. These larvae are submitted to transport and dispersion; they will feed in the plankton and will probably not experience the same conditions of the maternal environment, hypothetically reducing the necessity of energy transfer from mother to larvae. Thus, one could assume that the food environment experienced by mother would not impact the quality of the larvae produced. Interestingly, few authors have shown that, under stressful temperatures and low phytoplankton concentrations, maternal individuals of a tropical barnacle are able to manipulate the transfer of different types of fatty acids to their nauplii, a possible strategy to guarantee higher survival rates until this same nauplii encounters better food conditions in the water column [41]. Variations in the amount and type of phytoplankton available for planktotrophic larvae during development cycle interfere in the different larval traits, including in the success of metamorphosis into the juvenile stage. Larvae of gastropods [15, 16, 42], bivalves [36, 43, 44], and barnacles [45, 46] vary in size, development rate, and survival to the juvenile stage, in direct association with the quality and amount of phytoplankton offered them during their development.

Larvae must be able to survive from pelagic to benthic conditions and return to the rocky shore communities, in order to reach the adult phase. Settlement success and post-settlement survivorship are also matters of larval history [12, 15, 21], and many more. Contrary of what has been accepted for a long time, settlement of larvae in the benthic environment, and its metamorphose to the juvenile stage do not result in a "new beginning" for those individuals, but the feeding conditions experienced by larvae and its results on their physiological quality can be carried over to the next stage, and those individuals who faced low phytoplankton concentrations during its life in the plankton might become juveniles with lower growth and survival potential, influencing directly on the fate of that population [46–51].

#### **4. Larval transport, settlement, and recruitment**

Phytoplankton and larval abundances are sometimes controlled by the same oceanographic processes. Phytoplankton grows and reproduces under very specific environmental conditions, driven mainly by turbulence and nutrient availability [52]. Ocean movements, such as turbulence, vertical mixing, and currents, also affect larval abundance at small (e.g., Ref. [53]) and larger scales (e.g., Ref. [54]). Marine larvae take advantage of meso- and large-scale oceanographic features for transport and dispersion. These larvae have different responses depending on the velocity at that depth, assuming a specific swimming or orientation pattern (e.g., Ref. [55]). Besides, larvae are able to control their position in the water column and move together with the main current at that specific depth [56–58], what in turn might result in variability of larval supply in time and space [59]. Some oceanographic features that accumulate and transport marine invertebrate larvae are responsible for disturbing phytoplankton as well. For example, upwelling currents, which cause phytoplankton blooms by injecting cold nutrient-rich waters in the photic zone, may move larvae of rocky shore invertebrate to shallower waters (e.g., Refs. [60, 61]). Storms are other meteorological-oceanographic phenomena that disturb both chlorophyll-a concentration at the nearshore environments (e.g., Ref. [62]) and the larval abundances close to the rocky shores [63].

However, most invertebrates that inhabit the rocky shores produce planktotrophic larvae. These larvae are submitted to transport and dispersion; they will feed in the plankton and will probably not experience the same conditions of the maternal environment, hypothetically reducing the necessity of energy transfer from mother to larvae. Thus, one could assume that the food environment experienced by mother would not impact the quality of the larvae produced. Interestingly, few authors have shown that, under stressful temperatures and low phytoplankton concentrations, maternal individuals of a tropical barnacle are able to manipulate the transfer of different types of fatty acids to their nauplii, a possible strategy to guarantee higher survival rates until this same nauplii encounters better food conditions in the water column [41]. Variations in the amount and type of phytoplankton available for planktotrophic larvae during development cycle interfere in the different larval traits, including in the success of metamorphosis into the juvenile stage. Larvae of gastropods [15, 16, 42], bivalves [36, 43, 44], and barnacles [45, 46] vary in size, development rate, and survival to the juvenile stage, in direct association with the quality and amount of phytoplankton offered

Larvae must be able to survive from pelagic to benthic conditions and return to the rocky shore communities, in order to reach the adult phase. Settlement success and post-settlement survivorship are also matters of larval history [12, 15, 21], and many more. Contrary of what has been accepted for a long time, settlement of larvae in the benthic environment, and its metamorphose to the juvenile stage do not result in a "new beginning" for those individuals, but the feeding conditions experienced by larvae and its results on their physiological quality can be carried over to the next stage, and those individuals who faced low phytoplankton concentrations during its life in the plankton might become juveniles with lower growth and

Phytoplankton and larval abundances are sometimes controlled by the same oceanographic processes. Phytoplankton grows and reproduces under very specific environmental conditions, driven mainly by turbulence and nutrient availability [52]. Ocean movements, such as turbulence, vertical mixing, and currents, also affect larval abundance at small (e.g., Ref. [53]) and larger scales (e.g., Ref. [54]). Marine larvae take advantage of meso- and large-scale oceanographic features for transport and dispersion. These larvae have different responses depending on the velocity at that depth, assuming a specific swimming or orientation pattern (e.g., Ref. [55]). Besides, larvae are able to control their position in the water column and move together with the main current at that specific depth [56–58], what in turn might result in variability of larval supply in time and space [59]. Some oceanographic features that accumulate and transport marine invertebrate larvae are responsible for disturbing phytoplankton as well. For example, upwelling currents, which cause phytoplankton blooms by injecting cold nutrient-rich waters in the photic zone, may move larvae of rocky shore invertebrate to shallower waters (e.g., Refs. [60, 61]). Storms are other meteorological-oceanographic

survival potential, influencing directly on the fate of that population [46–51].

**4. Larval transport, settlement, and recruitment**

them during their development.

10 Chlorophyll

Settlement is a function of larval supply [64]. Consequently, successful settlement relies on larvae, which need to find suitable settlement sites and be able to metamorphose. In this phase of the larval cycle, biochemical and physical cues either stimulate or block settlement. The presence of biofilm on the rocks is very important for settling larvae, in particular for the sessile larvae, because biofilm may define if that is a favorable settlement spot. Biofilm characteristics control larval behavior during settlement [65]; as a result, settlement rates and the chlorophyll-a content in the biofilm are correlated [66]. Settlement may also be correlated with fluctuations in chlorophyll-a concentration just as a consequence of the coupling between phytoplankton blooms and larval release [12, 28]. When the latter situation is true, fluctuations in chlorophyll-a concentration and variations in settlement rates are time lagged in several days [35], what may depend on the time that the larva takes to fully develop. On the other hand, if larval supply and phytoplankton dynamics are controlled by the same features, as it was explained in the previous paragraph, peaks in chlorophyll-a concentrations and settlement rates will occur simultaneously (e.g., Ref. [20]).

Recruitment rates are regulated by fluctuations in the pelagic environment affecting larval supply [67]. Recruitment success means that settled larvae survived until they are able to reproduce. In the post-settlement period, phytoplankton availability in the benthos and pelagial can control the survivorship of settlers in rocky shore communities. Although most early recruits of rocky shore invertebrates are filter feeders, they do not have the same diet and they may be very selective [68], choosing determinate phytoplankton species as food items depending on their size. Changes in the phytoplankton community might benefit one or the other species depending on their feeding behavior [68]. Although the relationship between recruitment and chlorophyll-a concentration is influenced by species-specific characteristics, information on this subject is still relatively scarce for rocky shore invertebrates. Small- and large-scale spatial variability in recruitment of rocky shore invertebrates are related to local and regional gradients of chlorophyll-a concentration in the surface waters. Geographic barriers that restrict phytoplankton abundance are also responsible for setting geographical limits for recruitment at the rocky shores. Recruitment rates may vary in several orders of magnitude among regions and sites, potentially due to persistent gradients in phytoplankton availability, and in turn gradients in chlorophyll-a concentration (e.g., Refs. [69, 70]. Even sites within the same bay or just less than 1 km apart may present high contrasts in recruitment rates as a consequence of differences in the phytoplankton dynamics [71].

#### **5. The numerical relationships between chlorophyll-a concentration and larval processes**

Phytoplankton is a limiting resource to the survival of marine invertebrate larvae, as it was described throughout the chapter; consequently, chlorophyll-a concentration is a key factor regulating larval dynamics in rocky shore communities. Variations in larval processes and fluctuations in chlorophyll-a concentration tend to be highly correlated (e.g., trends of recruitment rates [69]). These correlations could be incorporated to ecological and numerical models to predict larval processes based on the values of chlorophyll-a concentration in the water (e.g., Ref. [72]). Although there are daily measurements of chlorophyll-a concentration in the ocean surfaces at a global scale, the levels of correlation between chlorophyll-a and larval dynamics are described only for a few species and some coastal areas.

Trends may be divided in groups according to the relationship between larval and phytoplankton dynamics. If the oceanographic processes promoting larval supply and settlement are also responsible for enabling phytoplankton growth and reproduction, variations in larval processes and in chlorophyll-a concentration may be positively correlated. On the other hand, if larval supply and settlement are enabled by less favorable conditions for phytoplankton, the fluctuations in the rates of larval processes may be negatively related to the concentrations of chlorophyll-a. Evidences of both trends were registered for rocky shore invertebrates in several regions [20, 21, 73]. Although the oceanographic and ecological processes that affect community dynamics are similar at the rocky shores, the correlation degrees between phytoplankton abundance and larval processes vary among sites and taxa. Correlations are stronger when reproduction and larval processes are regulated by the same mechanisms controlling phytoplankton blooms. For instance, in upwelling regions, these correlations are expected to be stronger [74], but may not be significant depending on the site (e.g., Ref. [75]). Barnacle and mussel recruits that occupy the same intertidal zone are not necessarily affected by fluctuations in chlorophyll-a concentration in similar ways, even presenting opposite trends in recruitment [21].

#### **6. Climate change on phytoplankton larval dynamics**

Climate change has important consequences for benthic-pelagic dynamics. Global warming has already caused alterations in the patterns of sea surface temperature and ocean currents, which in turn directly influenced the trends of phytoplankton abundance. Larvae and recruits of rocky shore invertebrates have to cope with such alterations in food availability concomitant to other climatic changes. The effects of phytoplankton and other climatic factors, such as water temperature, tend to be synergic [76]. Global warming conditions might not be positive for marine invertebrate larvae which, on one hand, survive under a wide range of conditions, but their fitness is highly influenced by changes in food availability. Short- and long-term consequences of climate change on phytoplankton larval dynamics were already detected for rocky shore communities. On the scale of decades, longer events of upwelling in the recent 20 years doubled the recruitment rates in some shores [77]. Results showed that, in small scale conditions, variability in phytoplankton has different effects on larval performance under different levels of climate change (Kasten, personal communication). However, how species will respond to multiple factors under *in situ* oceanic climatic conditions are hard to forecast, since information in larval dynamics are not available for most species and rocky shore systems.

### **7. Final considerations**

regulating larval dynamics in rocky shore communities. Variations in larval processes and fluctuations in chlorophyll-a concentration tend to be highly correlated (e.g., trends of recruitment rates [69]). These correlations could be incorporated to ecological and numerical models to predict larval processes based on the values of chlorophyll-a concentration in the water (e.g., Ref. [72]). Although there are daily measurements of chlorophyll-a concentration in the ocean surfaces at a global scale, the levels of correlation between chlorophyll-a and

Trends may be divided in groups according to the relationship between larval and phytoplankton dynamics. If the oceanographic processes promoting larval supply and settlement are also responsible for enabling phytoplankton growth and reproduction, variations in larval processes and in chlorophyll-a concentration may be positively correlated. On the other hand, if larval supply and settlement are enabled by less favorable conditions for phytoplankton, the fluctuations in the rates of larval processes may be negatively related to the concentrations of chlorophyll-a. Evidences of both trends were registered for rocky shore invertebrates in several regions [20, 21, 73]. Although the oceanographic and ecological processes that affect community dynamics are similar at the rocky shores, the correlation degrees between phytoplankton abundance and larval processes vary among sites and taxa. Correlations are stronger when reproduction and larval processes are regulated by the same mechanisms controlling phytoplankton blooms. For instance, in upwelling regions, these correlations are expected to be stronger [74], but may not be significant depending on the site (e.g., Ref. [75]). Barnacle and mussel recruits that occupy the same intertidal zone are not necessarily affected by fluctuations in chlorophyll-a concentration in similar ways, even presenting opposite

Climate change has important consequences for benthic-pelagic dynamics. Global warming has already caused alterations in the patterns of sea surface temperature and ocean currents, which in turn directly influenced the trends of phytoplankton abundance. Larvae and recruits of rocky shore invertebrates have to cope with such alterations in food availability concomitant to other climatic changes. The effects of phytoplankton and other climatic factors, such as water temperature, tend to be synergic [76]. Global warming conditions might not be positive for marine invertebrate larvae which, on one hand, survive under a wide range of conditions, but their fitness is highly influenced by changes in food availability. Short- and long-term consequences of climate change on phytoplankton larval dynamics were already detected for rocky shore communities. On the scale of decades, longer events of upwelling in the recent 20 years doubled the recruitment rates in some shores [77]. Results showed that, in small scale conditions, variability in phytoplankton has different effects on larval performance under different levels of climate change (Kasten, personal communication). However, how species will respond to multiple factors under *in situ* oceanic climatic conditions are hard to forecast, since information in larval dynamics are not available

larval dynamics are described only for a few species and some coastal areas.

**6. Climate change on phytoplankton larval dynamics**

trends in recruitment [21].

12 Chlorophyll

for most species and rocky shore systems.

Phytoplankton has a high regulatory potential in larval dynamics in the rocky shore communities. Rates of larval processes in rocky shore invertebrates are highly correlated with spatiotemporal fluctuations in chlorophyll-a concentration in the sea surfaces. The role of phytoplankton in larval dynamics at the community levels is not known, because information for most species is incipient. It is important to highlight that scientific improvements are needed to allow that use of variations chlorophyll-a concentration as a tool for modeling and forecasting the supply side ecology in rocky shore communities.

#### **Author details**

Ana Carolina de Azevedo Mazzuco1,2\* and Paula Kasten<sup>2</sup>

\*Address all correspondence to: ac.mazzuco@me.com

1 Federal University of Espírito Santo, Dept. of Oceanography, Vitória, ES, Brazil

2 Federal University of São Paulo, Institute of Ocean Sciences, Santos, SP, Brazil

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**Provisional chapter**

### **How Does Chloroplast Protect Chlorophyll Against Excessive Light? Excessive Light?**

**How Does Chloroplast Protect Chlorophyll Against** 

Lucia Guidi, Massimiliano Tattini and Marco Landi

Additional information is available at the end of the chapter Additional information is available at the end of the chapter

Lucia Guidi, Massimiliano Tattini and Marco

http://dx.doi.org/10.5772/67887

#### **Abstract**

Landi

Chlorophylls (Chls) are the most abundant plant pigments on Earth. Chls are located in the membrane of thylakoids where they constitute the two photosystems (PSII and PSI) of terrestrial plants, responsible for both light absorption and transduction of chemical energy via photosynthesis. The high efficiency of photosystems in terms of light absorption correlates with the need to protect themselves against absorption of excess light, a process that leads to the so-called photoinhibition. Dynamic photoinhibition consists of the downregulation of photosynthesis quantum yield and a series of photo-protective mechanisms aimed to reduce the amount of light reaching the chloroplast and/or to counteract the production of reactive oxygen species (ROS) that can be grouped in: (i) the first line of chloroplast defence: non-photochemical quenching (NPQ), that is, the dissipation of excess excitation light as heat, a process that takes place in the external antennae of PSII and in which other pigments, that is carotenoids, are directly involved; (ii) the second line of defence: enzymatic antioxidant and antioxidant molecules that scavenge the generated ROS; alternative electron transport (cyclic electron transport, pseudo-cyclic electron flow, chlororespiration and water-water cycle) can efficiently prevent the over-reduction of electron flow, and reduced ferredoxin (Fd) plays a key role in this context.

**Keywords:** antioxidant, carotenoids, excess excitation energy, non-photochemical quenching, photosystem

#### **1. Introduction**

Pigments in plants, cyanobacteria, algae and photosynthetic anoxygenic bacteria are the most important molecules involved in photosynthesis, the only biological process that tunnels

energy on Earth. Pigments play two key roles in photosynthesis: they absorb sunlight and transduce it into chemical energy. The most important pigment is certainly chlorophyll (Chl), an organic compound that typically shows chlorine, a cyclic tetrapyrrole ring, coordinated to a central atom of magnesium (**Figure 1**). This molecular structure is very similar to that found in the eme group in which the central atom is iron. Diversification of various Chls is due to the different side chains bonded to the chlorine ring (Chl *a*, *b*, *c*, *d*, *e* and *f*).

The process of light absorption consists of a sequence of photophysical and photochemical reactions that are subdivided into three stages: (i) light absorption, (ii) utilization of this energy to synthesize ATP and reducing power, reduced ferredoxin (Fd) and NADPH and (iii) absorption and reduction of atmospheric CO2 into carbon skeleton. However, the most important and true light reaction is represented by charge separation that occurs at the reaction

centres. The process is possible for the presence of organic molecules able to capture sunlight and transduce it in chemical energy namely photosynthetic pigments and that is chlorophylls and, carotenoids. These pigments aggregate with proteins and act as an antenna harvesting the energy of sunlight and tunnelling this energy into the reaction centres located in photosystems. In plants and algae, there are about 200–400 light harvesting molecules. Light harvesting complexes have evolved many adaptive mechanisms that permit photosynthetic organisms to thrive in different environments. The spectral distribution of sunlight that reaches our planet largely covers the absorption spectra of photosynthetic pigments utilized in light harvesting antennas (**Figure 2**). In a general way, light harvesting antennas have developed the ability to optimize light capture under both low- and high-intensity light conditions [1].

energy on Earth. Pigments play two key roles in photosynthesis: they absorb sunlight and transduce it into chemical energy. The most important pigment is certainly chlorophyll (Chl), an organic compound that typically shows chlorine, a cyclic tetrapyrrole ring, coordinated to a central atom of magnesium (**Figure 1**). This molecular structure is very similar to that found in the eme group in which the central atom is iron. Diversification of various Chls is due to the

The process of light absorption consists of a sequence of photophysical and photochemical reactions that are subdivided into three stages: (i) light absorption, (ii) utilization of this energy to synthesize ATP and reducing power, reduced ferredoxin (Fd) and NADPH and (iii)

tant and true light reaction is represented by charge separation that occurs at the reaction

into carbon skeleton. However, the most impor-

different side chains bonded to the chlorine ring (Chl *a*, *b*, *c*, *d*, *e* and *f*).

absorption and reduction of atmospheric CO2

22 Chlorophyll

**Figure 1.** Structures of the chlorophyll molecules.

The optimal absorption wavelength range for light harvesting antennas is in the red region (680–690 nm), where the energy is utilized by chlorophyll to split water and reduce ferredoxin. The evolution of the most abundant pigments, chlorophyll *a*, is probably related to its efficient absorption in this region in addition to, perhaps, its chemistry and for its redox potential.

All photosynthetic pigments show a chromophore, which possesses two orbitals whose difference in energy falls within the light spectrum. In consequence, a photon of incident light is able to excite an electron from its ground-state orbital to the excited state. From a chemical point of view, the chromophore exists as conjugated π-electron systems or metal complexes. In a conjugated π system, electron excitation occurs between π orbitals spread across alternating single and double bounds (e.g., carotenoids). The metal complex chromophores share d orbitals between transition metals and ligands (e.g., chlorophylls). Really, in the antenna pigments, chromophores are not individual entities, and they synergically interact with each other and this interaction plays a crucial role in the light harvesting mechanism.

**Figure 2.** Chlorophyll *a*, *b* and carotenoids absorbance spectra.

Light-harvesting complex (LHC) is the complex of subunit proteins that may be part of a larger supercomplex of a photosystem and is the functional unit in photosynthesis, devoted to the absorption of sunlight. The energy excitation is first tunnelled among other surrounding molecules of the same complex and then from one LHC to another and then funnelled to reaction centres (RCs), where it is converted into charge separation with 90% quantum efficiency.

The presence of proteins in LHC complexes is attributable to the fact that Chl of RCs cannot absorb sunlight at an efficient rate that is enough for efficient photosynthesis to occur. In fact, Chl molecules in RCs absorb only a few photons each second, which are insufficient to drive electron transport into chloroplast membranes (present in 1 RC of about 300 antenna molecules). To overcome this problem, RCs are associated with antenna pigment-protein complexes that absorb sunlight and very efficiently transfer it to RCs. For the importance of the LHCs in gathering sunlight, they differ in the number of pigments and in their composition and structure in a way that they are an optimized energy collector system (**Figure 3**). The proteins play an important function in the precise position, mutual separation and relative orientation of antenna.

**Figure 3.** A schematic representation of the light absorption process of chloroplasts. Antenna complexes, composed of carotenoids, Chl *a* and Chl *b* molecules, absorb photons from sunlight and transfer them to the RC, which consists of a special couple of molecules of Chl *a*. Antenna complexes and the RC form a photosystem.

Photosynthetic unit (PSU) represents the basic unit of the light-harvesting apparatus and consists of a large number of antenna chromophores coupled to a RC. Excitation-transfer pathways follow a scheme in which different chromophores build an energy funnel where chromophores, which absorb in the blue side of spectrum, transfer excitation energy to more red-shifted chromophores (**Figure 3**). Theoretically, the PSUs are considered individual entities but [2] proposed the *lake* and *puddle* model. In the second model, the PSUs do not interact with each other and the excitation light absorbed by chromophores is always transferred to the same RC. Differently, in the *lake* model, the antenna chromophores form a matrix with embedded RCs in which there is an unrestricted energy transfer.

#### **2. Charge separation in photosystems and electron transport**

Light-harvesting complex (LHC) is the complex of subunit proteins that may be part of a larger supercomplex of a photosystem and is the functional unit in photosynthesis, devoted to the absorption of sunlight. The energy excitation is first tunnelled among other surrounding molecules of the same complex and then from one LHC to another and then funnelled to reaction centres (RCs), where it is converted into charge separation with 90% quantum

The presence of proteins in LHC complexes is attributable to the fact that Chl of RCs cannot absorb sunlight at an efficient rate that is enough for efficient photosynthesis to occur. In fact, Chl molecules in RCs absorb only a few photons each second, which are insufficient to drive electron transport into chloroplast membranes (present in 1 RC of about 300 antenna molecules). To overcome this problem, RCs are associated with antenna pigment-protein complexes that absorb sunlight and very efficiently transfer it to RCs. For the importance of the LHCs in gathering sunlight, they differ in the number of pigments and in their composition and structure in a way that they are an optimized energy collector system (**Figure 3**). The proteins play an important function in the precise position, mutual separation and relative

**Figure 3.** A schematic representation of the light absorption process of chloroplasts. Antenna complexes, composed of carotenoids, Chl *a* and Chl *b* molecules, absorb photons from sunlight and transfer them to the RC, which consists of a

special couple of molecules of Chl *a*. Antenna complexes and the RC form a photosystem.

efficiency.

24 Chlorophyll

orientation of antenna.

Photosynthesis starts with light absorption by the chromophores, which excites the molecules from the ground state to an electronic excited state. Once sunlight energy is absorbed, pigments in the excited state have a short life and relax to the ground state after about 4 ns [3]. The singlet excited state lifetime of Chl is lower compared with the radiative lifetime, largely owing to intersystem crossing, which yields triplet excited states of Chl (about 10 ns) [4]. This electronic excitation must be usefully harvested before the molecules relax, and this happens when excitons are transferred through space among chromophores until they reach, eventually, a RC where charge separation occurs. In plants, there are two RCs constituted by two Chl molecules, P680 and P700, respectively, for PSII and PSI, and Chl with absorbance maxima corresponding to these wavelengths is proposed as the final slight sink. These chlorophylls drive electron transfer by charge separation, a reaction in which P680 and P700 molecules reduce an acceptor. These driving reactions energetically downhill from the potential that is more negative to ones that are more positive (**Figure 4**). All these electron transfer steps in photosynthesis share a common feature. The loss of an electron from one component, which remains in an oxidized state, reduces another one. Typically, electron transport carriers are small molecules or atoms of metallic elements that can exist in a number of valence states.

In photosystem RCs, the light-induced loss of an electron (charge separation) leaves P680 and P700 in an oxidized state (P680+ and P700+ ) and the respective acceptors, pheophytin for P680 and A0 (chlorophyll), in a reduced state. P680+ is reduced from an adjacent tyrosine molecule (TyrZ) in the polypeptide chain of the D1 protein of the PSII complex. In turn, the oxidized is reduced by electrons from the oxygen-evolution complex (OEC) that oxidized water. Two water molecules are oxidized to produce oxygen, four protons and four electrons that are transferred one at a time. These redox reactions are carried out by OEC that consists of four manganese atoms held in a protein matrix with one atom of calcium and chlorine each (**Figure 4**). This process is known as a S-cycle from [5] that provides protons derived from water oxidation to be released into the lumen of the thylakoid membranes.

In the other set of reactions, reduced pheophytin is oxidized by passing an electron to the first of two plastoquinone (PQ) molecules, tightly bound at the site QA of D2 protein in the PSII. Then, via an iron atom, an electron is transferred to the next PQ at the site QB. Both PQs require two electrons for their complete reduction; at the QA site, PQ undergoes to a single

**Figure 4.** A representation of the linear non-cyclic (solid line) and cyclic electron flow (dashed line) in the chloroplast membranes. OEC tetranuclear Mn cluster; P680, reaction centre of photosystem II (PSII); P680\*, excited electronic state of P680; Ph, pheophytin; QA and QB, plastoquinone; protein complex containing cytochrome b6 and cytochrome *f*; PC, plastocyanin; P700, reaction centre of PSI; P700\*, excited electronic state of P700; A0 , a special chlorophyll *a* molecule; A1 , phylloquinone; Fe-S, iron sulphur centres; Fd, ferredoxin; NADP, nicotinamide-adenine dinucleotide phosphate and FNR, ferredoxin-NADP<sup>+</sup> reductase.

reduction event to the semiquinone state before being re-oxidized by the PQ at QB site. Two successive reductions occur that fully reduce PQ at QB site, which, for its reduction, requires also two protons from the stromal side of the membranes and forms PQH<sup>2</sup> that leaves PSII and diffuses in the lipid bilayer, representing a mobile carrier of protons and electrons. A new molecule of PQ (in oxidized form) replaces this plastoquinone in the QB site.

PQH<sup>2</sup> formed by the PSI activity represents the substrate of the Q cycle on cytochrome *b6 f*, another integral transmembrane protein complex on thylakoid membranes. PQH<sup>2</sup> is oxidized in two steps to PQ. The first step happens at Qp site, located on the luminal side of cytochrome *b6 f*, and the electron is transferred at the end to plastocyanin (PC), a soluble small protein containing copper. The second electron is transferred until Qn site located on the stromal side of the cytochrome where it reduces further PQ molecule to semiplastoquinone. Another PQH<sup>2</sup> molecule originating from PSII is oxidized in the same two steps at the Qp site, generating further a reduced plastocyanin and completing the reduction of semiplastoquinone to PQH<sup>2</sup> . The oxidation of PQH<sup>2</sup> at Qp site determines the release of two protons in the lumen that represents the most important feature of the Q cycle. In fact, this cycle acts as a proton pump, essential to generate the transmembrane electrochemical H<sup>+</sup> gradient.

After light absorption and charge separation in PSI, P700+ is generated, and it is reduced back to P700 by direct interaction with reduced PC diffusing from cytochrome *b6 f* complex. Plastocyanin, from its copper atom, reduces directly P700+ . The electron flow generated by charge separation that occurs in P700 determines the reduction of different carriers, and the final electron acceptor is represented by Fd, a small water-soluble iron-sulphur protein. Reduced Fd is capable of reducing a variety of molecules. Usually, it reduces NADP<sup>+</sup> , which requires two electrons and two protons to yield NADPH in a reaction catalyzed by ferredoxin-NADP+ reductase (FNR) (**Figure 4**), thus completing the so-called Z scheme. The electron flow generates even chemical energy, that is ATP, by the enzymatic activity of ATP-ase, a transmembrane complex that utilizes the proton gradient generated by Q cycle and water oxidation, to synthetize ATP.

#### **3. Excess of excitation energy**

reduction event to the semiquinone state before being re-oxidized by the PQ at QB site. Two successive reductions occur that fully reduce PQ at QB site, which, for its reduction, requires

, phylloquinone; Fe-S, iron sulphur centres; Fd, ferredoxin; NADP, nicotinamide-adenine dinucleotide phosphate and

**Figure 4.** A representation of the linear non-cyclic (solid line) and cyclic electron flow (dashed line) in the chloroplast membranes. OEC tetranuclear Mn cluster; P680, reaction centre of photosystem II (PSII); P680\*, excited electronic state

and diffuses in the lipid bilayer, representing a mobile carrier of protons and electrons. A new

in two steps to PQ. The first step happens at Qp site, located on the luminal side of cytochrome *b6 f*, and the electron is transferred at the end to plastocyanin (PC), a soluble small protein containing copper. The second electron is transferred until Qn site located on the stromal side of the cytochrome where it reduces further PQ molecule to semiplastoquinone. Another PQH<sup>2</sup> molecule originating from PSII is oxidized in the same two steps at the Qp site, generating further a reduced plastocyanin and completing the reduction of semiplastoquinone to PQH<sup>2</sup>

represents the most important feature of the Q cycle. In fact, this cycle acts as a proton pump,

back to P700 by direct interaction with reduced PC diffusing from cytochrome *b6 f* complex.

formed by the PSI activity represents the substrate of the Q cycle on cytochrome *b6 f*,

at Qp site determines the release of two protons in the lumen that

gradient.

is generated, and it is reduced

that leaves PSII

and cytochrome *f*; PC,

, a special chlorophyll *a* molecule;

is oxidized

.

also two protons from the stromal side of the membranes and forms PQH<sup>2</sup>

of P680; Ph, pheophytin; QA and QB, plastoquinone; protein complex containing cytochrome b6

plastocyanin; P700, reaction centre of PSI; P700\*, excited electronic state of P700; A0

reductase.

molecule of PQ (in oxidized form) replaces this plastoquinone in the QB site.

essential to generate the transmembrane electrochemical H<sup>+</sup>

After light absorption and charge separation in PSI, P700+

another integral transmembrane protein complex on thylakoid membranes. PQH<sup>2</sup>

PQH<sup>2</sup>

A1

26 Chlorophyll

FNR, ferredoxin-NADP<sup>+</sup>

The oxidation of PQH<sup>2</sup>

In the past, the higher order structure of PSII was thought to be important only to increase the efficiency of light harvesting; nowadays, it has been suggested that it provides the essential dynamic properties involved in its regulation [6]. When light is low, in a way, extremely efficient antenna systems absorb light and tunnel it through RC, but when light is in excess, a large extent of this energy is dissipated, overall as heat, to prevent photo-damage to PSUs. When plants are exposed to shade or sunlight conditions, different mechanisms occur. Shade leaves are typically larger in area but thinner than sun leaves because they develop shorter palisade cells. In shade leaves, the chloroplasts move within the cells to take up a position where they will absorb the maximum light without shading other chloroplasts below. In addition, shade leaves show a large number of antenna, and usually, the peripheral antenna are rich in Chl *b* molecules (Chl *a*/*b* = 1.33). All these mechanisms enhance and optimize the light absorption. However, even shade leaves have adapted mechanisms aimed to regulate the light absorption, as the state II-I transition (also called spillover process). The aim of this process is the reduction of light tunnelled to P680 altering the ratio of light energy absorbed between PSII and PSI. In fact, RCs of the two photosystems have different absorption spectra (high energy is absorbed by P680 as compared with P700), and this determines that when the energy flow through each is not balanced to the requirement of the Z scheme, an excess of energy could accumulate in the system. In this way, LHCII trimers represent a feedback loop that adjusts the amount of antenna Chls, providing energy to each photosystem (state transition). The excess of light energy flowing through PSII RCs is higher than that flowing through PSI RCs, conditions in which an excess of reduced PQ occurs. This activates a kinase that phosphorylates some LHCII trimers, and this extra charge allows them to dissociate from the PSII (state II) and migrate towards the stroma lamellae (state I transition) where they bind to the PSI complex, increasing in this way the flow through the system. The increase of PSI activity leads to the oxidation of reduced PQ, which activates a phosphatase that removes the phosphate group to the LHCII trimers that return to PSII (state II transition).

In contrast, sun leaves live in very high radiation levels overall at the top of the canopy. The light response curve in relation to the light intensity shows that the amount of energy utilized is lower than that absorbed because the light energy utilized in carbon reduction is mostly due to the limitation on the rate of CO2 diffusing into the leaf (**Figure 5**). In these conditions,

**Figure 5.** Absorbed and utilized energy in response to increasing light intensities. When light absorbed exceed photosystems requirement, the 'excess energy' can potentially cause photo-oxidative damage if it is not efficiently dissipated.

the antenna Chls become saturated and tunnel a high flow of the excitation energy to the RC that cannot be dissipated along the electron flow. The excess of energy must be efficiently dissipated through different mechanisms in order to avoid photo-damage to PSII.

Photosystem II is particularly sensitive to photoinhibition because the high redox potential of the oxidized P680 (P680+ ), on the other hand, necessary for water oxidation. Accumulation of P680+ leads to different types of photoinhibition:


Different mechanisms are present in PSII aimed to dissipate the excess of photons absorbed by antenna, and different defence lines occur into the chloroplast.

#### **4. First line of defence of chloroplast: dissipation of excess excitation light**

First line of chloroplast defence includes suppression mechanisms aimed to reduce or dissipate the excitation light tunnelled in P680. At leaf level, the change in the leaf angle with respect to the incident light and/or the chloroplast movement into the leaf to self-shading positions along the sidewalls of cells represent mechanisms by which a decrease in absorbed light can occur.

In the chloroplast, there are essentially three mechanisms to contrast the high light conditions: adjustment in synthesis and amount of antenna protein, movement of LHCII (state II-I transition) and non-photochemical quenching [7]. The first of these mechanisms is related to the expression of *Lhcb* genes, whose expression is downregulated by high light conditions and/or low CO2 concentration. The sensor mechanism is not known even though one possible candidate is the redox potential (i.e., the level of reduced PQ) [8], but also ROS represent possible signal molecules [9, 10]. Clearly, these slow mechanisms cannot entirely prevent the accumulation of excess of energy in the antenna system. However, photosynthesis in green plants depends on protective mechanisms that adapt within minutes or seconds to changing light conditions. Excited Chls return to the ground state either by emitting photons (fluorescence) or by dissipating it as heat. All these mechanisms aimed to remove this trapped energy before it passed on down the electron transport chain are named *non-photochemical quenching* (NPQ). NPQ is heterogeneous and composed by at least three components: the major and rapid component is the pH- or energy-dependent component qE, a second component qT, related to the phenomenon of state transition but negligible in most of plants under excess light and the third and slow component, qI, related to the photoinhibition of photosynthesis [11].

It has been reported that two distinct qE mechanisms occur, one involving zeaxanthin (Zea) (quenching type 1) and the other carotenoid lutein (Lut) (quenching type 2) [12]. In qE type I, three xanthophylls, violaxanthin (Vio), anteraxanthin (Ant) and Zea, are involved in the well-known xanthophyll cycle in which the epoxidation of Vio to Zea via Ant determines an efficient dissipation of excess light into heat [13]. Electron flow pumping and generating protons in the lumen decrease its pH from about 7 to less than 5; this represents a strong signal that starts a series of quenching processes. The low pH-induced protonation of PsbS peptide, for its proximity to antenna complexes (CP24, CP26 and CP29), induces in turn in these complexes conformational changes. In the chemical state, antenna complexes bind one molecule of Zea and one of Chl (*Zea-Chl complex = quenching complex*) that accept energy transfer from excited Chls. Zeaxanthins are able to return to their ground state dissipating energy as heat :

the antenna Chls become saturated and tunnel a high flow of the excitation energy to the RC that cannot be dissipated along the electron flow. The excess of energy must be efficiently dis-

**Figure 5.** Absorbed and utilized energy in response to increasing light intensities. When light absorbed exceed photosystems requirement, the 'excess energy' can potentially cause photo-oxidative damage if it is not efficiently dissipated.

Photosystem II is particularly sensitive to photoinhibition because the high redox potential of

recombination is inhibited and P680 is expected to lead to the triplet state of P680, TP680\*. This chemical species may react with oxygen and produce harmful singlet oxygen. (ii) Donor-side photoinhibition: if the OEC is chemically inactivated, the donation of electrons from water does not keep up with the electron transfer from P680 to the accep-

chemical species induces the oxidation of various organic components such as proteins

(i) Acceptor-side photoinhibition: when reduced PQ is not re-oxidized, the P680\*

), on the other hand, necessary for water oxidation. Accumulation of

occurs. The high redox potential of this

charge

sipated through different mechanisms in order to avoid photo-damage to PSII.

leads to different types of photoinhibition:

tor side. In this case, an accumulation of P680+

or pigments until damage is done to D1 protein of PSII.

the oxidized P680 (P680+

P680+

28 Chlorophyll

LHCII\* + zeaxanthin → LHCII + zeaxanthin \* . (1)

Zeaxanthin \* → zeaxanthin + heat. (2)

It has been reported that in the crystal structure of LHCII is present Vio, and its peripheral localization suggests that it could be de-epoxidized to Zea by Vio de-epoxidase (VDE), an enzyme that is activated by low lumen pH occurring in high light conditions. The back reaction by Zea epoxidase is slow and causes a sustained quenching that relaxes within 1–3 hours following light stress and depends on the release of Zea from antenna pigments. In conclusion, Zea is certainly considered a regulator of light harvesting for its role in the xanthophyll cycle and carries out three fundamental roles during high light conditions: (i) protection against photo-oxidation due to radical oxygen's attack (because it quenches oxygen singlet energy), (ii) absorption of Chl triplet energy and (iii) absorption of incoming photons and transferring them to neighbouring Chl molecules increasing in this way the overall absorption spectrum of the PSs [14]. In addition, it has been reported that this xanthophyll exhibits an antioxidant function in the thylakoid membrane [15].

In addition, trimeric LHCII binds other types of xanthophylls: two all-*trans*-luteins and a 9-*cis*noexanthin [16]. The minor monomeric complexes CP24, CP26 and CP29 all bind Lut, and in addition, CP29 binds two xanthophyll cycle carotenoids and one-half to one neoxanthin (Neo), CP24 binds two xanthophyll cycle carotenoids and CP26 binds one xanthophyll cycle carotenoids and one Neo [17, 18]. In the quenching type 2, qE is an intrinsic LCHII property: protein conformational changes alter configurations of bound pigment (normally Lut), which become an efficient quencher of Chl-excited state [12]. A change in the conformational state of another LHCII-bound xanthophyll, Neo, correlates with the extent of quenching. In the model for type 2 quenching proposed by [19], Zea acts not as a quencher but as an allosteric modulator of the ΔpH sensitivity of this intrinsic LHCII quenching process. The two types of quenching involved different xanthophylls that operate at different sites, but there are some similarities in the reasons that both involve ΔpH and PsbS-mediated conformational changes [12].

Given that the xanthophyll cycle quenches only 95% of the triplet Chl [20], the unquenched triplet Chl is the reason for the need of singlet oxygen not only scavenging by carotenoids bound to LHCII but also by carotenoids free in lipid matrix [21]. Lut has the specific property of quenching harmful 3Chl\* by binding at site L1 of the major LHCII complex and of other Lhc proteins of plants, thus preventing ROS formation [20]. Neo contributes PSII photoprotection in a dual way: determins conformational change in trimeric LHCII, which reduces light absorption and controls the accessibility of the O2 to the inner core of the complex [20, 22]. The trimeric organization of LHCII is, definitively, effective in screening the internal protein domain from molecular oxygen [23].

#### **5. Second line of defence of chloroplast: antioxidant enzymes and molecules**

As reported above, the excess of excitation energy induces an excess of singlet-excited Chl *a* that is de-excited via thermal dissipation. However, the remaining singlet-excited Chl *a* can convert to triplet-excited Chl that readily reduces molecular oxygen. This determines the synthesis of ROS that is potentially dangerous to organic molecules in the chloroplast. In the second line of defence, antioxidant molecules and enzymes that together scavenge ROS play a key role.

The primary products of molecular oxygen reduction are disproportionate to H<sup>2</sup> O2 and O2 in a reaction catalyzed by superoxide dismutase (SOD). H<sup>2</sup> O2 produced is then reduced to water with the reducing power of ascorbate (ASA) in a reaction catalyzed by ASA peroxidase (APX), and ASA is oxidized to monodehydroascorbate (MDHA) that is directly reduced to ASA by reduced ferredoxin or NADPH by MDHA reductase. Alternatively, MDHA is spontaneously disproportionated to dehydroascorbate (DHA) and ASA. DHA is then reduced by reduced glutathione (GSH), by the enzyme DHA reductase that produces oxidized glutathione (GSSG) and ASA. Finally, GSSG is reduced again in GSH by the action of GSH reductase, and the reducing power is represented by reduced Fd or NADPH, that, in turn, are reduced by PSI activity. This indicates that any pathway aimed to regenerate ASA utilizes electrons derived from water. For this reason, the previous process is referred as water-water cycle [10].

It has been reported that in the crystal structure of LHCII is present Vio, and its peripheral localization suggests that it could be de-epoxidized to Zea by Vio de-epoxidase (VDE), an enzyme that is activated by low lumen pH occurring in high light conditions. The back reaction by Zea epoxidase is slow and causes a sustained quenching that relaxes within 1–3 hours following light stress and depends on the release of Zea from antenna pigments. In conclusion, Zea is certainly considered a regulator of light harvesting for its role in the xanthophyll cycle and carries out three fundamental roles during high light conditions: (i) protection against photo-oxidation due to radical oxygen's attack (because it quenches oxygen singlet energy), (ii) absorption of Chl triplet energy and (iii) absorption of incoming photons and transferring them to neighbouring Chl molecules increasing in this way the overall absorption spectrum of the PSs [14]. In addition, it has been reported that this xanthophyll exhibits an

In addition, trimeric LHCII binds other types of xanthophylls: two all-*trans*-luteins and a 9-*cis*noexanthin [16]. The minor monomeric complexes CP24, CP26 and CP29 all bind Lut, and in addition, CP29 binds two xanthophyll cycle carotenoids and one-half to one neoxanthin (Neo), CP24 binds two xanthophyll cycle carotenoids and CP26 binds one xanthophyll cycle carotenoids and one Neo [17, 18]. In the quenching type 2, qE is an intrinsic LCHII property: protein conformational changes alter configurations of bound pigment (normally Lut), which become an efficient quencher of Chl-excited state [12]. A change in the conformational state of another LHCII-bound xanthophyll, Neo, correlates with the extent of quenching. In the model for type 2 quenching proposed by [19], Zea acts not as a quencher but as an allosteric modulator of the ΔpH sensitivity of this intrinsic LHCII quenching process. The two types of quenching involved different xanthophylls that operate at different sites, but there are some similarities in

the reasons that both involve ΔpH and PsbS-mediated conformational changes [12].

Given that the xanthophyll cycle quenches only 95% of the triplet Chl [20], the unquenched triplet Chl is the reason for the need of singlet oxygen not only scavenging by carotenoids bound to LHCII but also by carotenoids free in lipid matrix [21]. Lut has the specific property of quenching harmful 3Chl\* by binding at site L1 of the major LHCII complex and of other Lhc proteins of plants, thus preventing ROS formation [20]. Neo contributes PSII photoprotection in a dual way: determins conformational change in trimeric LHCII, which reduces light

The trimeric organization of LHCII is, definitively, effective in screening the internal protein

As reported above, the excess of excitation energy induces an excess of singlet-excited Chl *a* that is de-excited via thermal dissipation. However, the remaining singlet-excited Chl *a* can convert to triplet-excited Chl that readily reduces molecular oxygen. This determines the synthesis of ROS that is potentially dangerous to organic molecules in the chloroplast. In the second line of defence, antioxidant molecules and enzymes that together scavenge ROS play a key role.

**5. Second line of defence of chloroplast: antioxidant enzymes and** 

to the inner core of the complex [20, 22].

antioxidant function in the thylakoid membrane [15].

absorption and controls the accessibility of the O2

domain from molecular oxygen [23].

**molecules**

30 Chlorophyll

In addition to the primary antioxidant systems, carotenoids have a protective role against ROS since they are very efficient physical and chemical quenchers of singlet oxygen and potent scavengers of other free radicals [24]. For example, β-carotene, located in the core complex of both PSII and PSI, plays a role as a quencher of Chl triplet and singlet oxygen [25], and the products generated from the oxidation of β-carotene by singlet oxygen represent primary sensor signalling under oxidative stress [26]. Other carotenoids play an important role as antioxidants in the chloroplast. Lut is the most abundant carotenoid in the chloroplast and is required as a quencher [7], while Neo can scavenge superoxide anion [27]. The antioxidant activity of carotenoids is carried out in combination with other lipophilic antioxidants. In this way, it has been reported that Zea, in cooperation with tocopherol, prevented photooxidation induced by high light [28], or a strong increase in carotenoids pigment (including those involved in xanthophyll cycle) is reported together with the activity of SOD enzyme following oxidative stress [29]. Again, carotenoids can influence the structure and fluidity of thylakoid membranes [30], that is essential for photosynthetic functions, influence barrier status to ions and oxygen, increase thermostability and protect against lipid peroxidation. In fact, as reported by [30], β-carotene can fluidize the membrane because it can move in the inner hydrophobic part of the membrane, and xanthophyll (and in particular Zea) shows the polar group that orientates these carotenoids perpendicular to the membrane surface.

#### **6. From PSII repair processes to alternative electron sinks**

In the last 30–40 years, the susceptibility of D1 protein to photo-damage has been well known, and the concept of the replacement of the damaged D1 protein during the repair cycle of PSII is extensively investigated [13, 31–33]. Moreover, D1 damage has been shown to be directly proportional to light intensity [34].

The repair process of photo-damaged D1 proteins consists of different steps: (i) prompt, partial disassembly of the PSII holocomplex, (ii) exposure of the photo-damaged PSII core to the stroma of the chloroplast, (iii) degradation of photo-damaged D1, (iv) *de novo* D1 biosynthesis and insertion in the thylakoid membrane and (v) re-assembly of the PSII holocomplex, followed by activation of the electron-transport process through the reconstituted D1/D2 heterodimer [35]. The sequence leading to the recovery of photo-damaged PSII is consistent with the frequent D1 turnover in the chloroplast and with the heterogeneity in the configuration and function of PSII.

In the past, the sensibility of PSII was linked to an inherent defect of photosynthetic apparatus but now it is clear how this mechanism of damage-repair of PSII is extremely regulated [33] and protects even PSI from irreversible damage. In fact, the repair mechanisms in PSI are time and high energy consuming, and it has been suggested that the inhibition of PSII is likely to protect PSI [33].

Reduced Fd plays an important role in preventing the over-reduction of electron flow, and a wide range of electron sinks are available in chloroplasts. Electrons are preferentially utilized by the FNR enzyme that produces NADPH for CO<sup>2</sup> photoassimilation or ferredoxin:thioredoxin reductase that synthesizes thioredoxin responsible for the regulation of some enzymes of Calvin-Benson cycle [36]. On the other hand, reduced Fd can release electrons also to ferredoxin:nitrite reductase and sulphite reductase for the reductive assimilation of nitrite [37] and sulphur [38]. Finally, reduced Fd represents an electron donor for fatty acid desaturases [39] and glutamine:oxoglutarate amino transferase [40]. However, when NADP<sup>+</sup> is not available, reduced Fd releases its electron to different acceptors whose function is to avoid an overreduction of PSI [41]. It has been discovered that there is an electron transport driven solely by PSI and scientists called it cyclic electron flow. In this cycle, electrons can be recycled from reduced Fd to PQ and subsequently, to the cytochrome *b6 f* complex via the Q cycle [42]. Such cyclic flow generates ΔpH and thus ATP without the accumulation of reduced species. In addition, the generated ΔpH may regulate photosynthesis via NPQ (see Section 4). Another electron acceptor of reduced Fd is molecular oxygen inducing the pseudo-cyclic electron flow. The reduction of molecular oxygen with one electron generates superoxide anions in the so-called Mehler reaction, which restores the redox poise when linear electron flow is over-reduced [43]. The radical oxygen species is efficiently removed by water-water cycle. Chlororespiration is another effective electron sink in which reduced Fd is directly involved. In this process, two enzymes play the key role: NADH dehydrogenase complex and nucleus-encoded plastidlocalized terminal oxidase (PTOX). The enzyme PTOX catalyzes the reaction in which electrons are transferred from PQH<sup>2</sup> to molecular oxygen forming water [44].

Finally, in addition to the above-reported electron flow, photorespiration is another efficient pathway by which plants adjust the ATP/NADPH ratio and consume the excess of excitation energy.

#### **7. Conclusions**

Certainly, Chls represent the key molecules involved in light energy absorption and transduction into chemical energy. Chls absorb the light energy that reaches leaves in a very efficient manner but sometimes, light exceeds photochemistry requirement, and the complexity of photosystems is essential to modulate and dissipate excess of excitation energy. A wide range of responses to environmental stimuli thus characterizes the photoprotection of chloroplasts. The increasing level of complexity from the molecular (pigments and protein) to supramolecular (photosystems) level mirrors the necessity of different time-scale responses (from seconds to months) to modulate light that is (inevitably) absorbed. In the range of seconds to minutes, modulation of the redox state of photosynthetic electron transport activates the non-photochemical quenching of excess of excitation energy not only through xanthophyll cycles [13] but also by II-I state transition [45]. On a larger scale (minutes to hours), modulation of redox state of electron transport induces changes in gene expression (organellar and nucleus) through retrograde regulation that changes the structure of the photosynthetic apparatus [46, 47]. On the time scale from weeks to months, the redox state of electron transport determines changes in plant growth and morphology [48].

#### **Author details**

the frequent D1 turnover in the chloroplast and with the heterogeneity in the configuration

In the past, the sensibility of PSII was linked to an inherent defect of photosynthetic apparatus but now it is clear how this mechanism of damage-repair of PSII is extremely regulated [33] and protects even PSI from irreversible damage. In fact, the repair mechanisms in PSI are time and high energy consuming, and it has been suggested that the inhibition of PSII is likely to

Reduced Fd plays an important role in preventing the over-reduction of electron flow, and a wide range of electron sinks are available in chloroplasts. Electrons are preferentially utilized by

reductase that synthesizes thioredoxin responsible for the regulation of some enzymes of Calvin-Benson cycle [36]. On the other hand, reduced Fd can release electrons also to ferredoxin:nitrite reductase and sulphite reductase for the reductive assimilation of nitrite [37] and sulphur [38]. Finally, reduced Fd represents an electron donor for fatty acid desaturases

able, reduced Fd releases its electron to different acceptors whose function is to avoid an overreduction of PSI [41]. It has been discovered that there is an electron transport driven solely by PSI and scientists called it cyclic electron flow. In this cycle, electrons can be recycled from reduced Fd to PQ and subsequently, to the cytochrome *b6 f* complex via the Q cycle [42]. Such cyclic flow generates ΔpH and thus ATP without the accumulation of reduced species. In addition, the generated ΔpH may regulate photosynthesis via NPQ (see Section 4). Another electron acceptor of reduced Fd is molecular oxygen inducing the pseudo-cyclic electron flow. The reduction of molecular oxygen with one electron generates superoxide anions in the so-called Mehler reaction, which restores the redox poise when linear electron flow is over-reduced [43]. The radical oxygen species is efficiently removed by water-water cycle. Chlororespiration is another effective electron sink in which reduced Fd is directly involved. In this process, two enzymes play the key role: NADH dehydrogenase complex and nucleus-encoded plastidlocalized terminal oxidase (PTOX). The enzyme PTOX catalyzes the reaction in which elec-

to molecular oxygen forming water [44].

Finally, in addition to the above-reported electron flow, photorespiration is another efficient pathway by which plants adjust the ATP/NADPH ratio and consume the excess of excitation

Certainly, Chls represent the key molecules involved in light energy absorption and transduction into chemical energy. Chls absorb the light energy that reaches leaves in a very efficient manner but sometimes, light exceeds photochemistry requirement, and the complexity of photosystems is essential to modulate and dissipate excess of excitation energy. A wide range of responses to environmental stimuli thus characterizes the photoprotection of chloroplasts. The increasing level of complexity from the molecular (pigments and protein) to supramolecular

[39] and glutamine:oxoglutarate amino transferase [40]. However, when NADP<sup>+</sup>

photoassimilation or ferredoxin:thioredoxin

is not avail-

and function of PSII.

32 Chlorophyll

protect PSI [33].

the FNR enzyme that produces NADPH for CO<sup>2</sup>

trons are transferred from PQH<sup>2</sup>

energy.

**7. Conclusions**

Lucia Guidi1 \*, Massimiliano Tattini<sup>2</sup> and Marco Landi1

\*Address all correspondence to: lucia.guidi@unipi.it

1 Department of Agriculture, Food and Environment, University of Pisa, Pisa, Italy

2 National Research Council of Italy, Department of Biology, Agriculture and Food Sciences, Institute for Sustainable Plant Protection, Sesto Fiorentino, Florence, Italy

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#### **Effects on the Photosynthetic Activity of Algae after Exposure to Various Organic and Inorganic Pollutants: Review Exposure to Various Organic and Inorganic Pollutants: Review**

**Effects on the Photosynthetic Activity of Algae after** 

DOI: 10.5772/67991

Andreas S. Petsas and Maria C. Vagi Andreas S. Petsas and Maria C. Vagi Additional information is available at the end of the chapter

Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/67991

#### **Abstract**

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Algal studies remain necessary for risk assessment and their utility in ecotoxicology is the evaluation of lethal and sub-lethal toxic effects of potential toxicants on inhabitants of several ecosystems. Effects on algal photosynthetic apparatus caused by various chemi‐ cal species have been extensively studied. The present chapter summarizes the pub‐ lished data concerning the toxicity of various organic and inorganic pollutants such as oils, pesticides, antifoulants and metals on photosynthesis of aquatic primary producers. Biochemical mode of action resulting in the disruption of photosynthesis depends on the chemical's nature and the characteristics of the exposed microorganism. Observed differ‐ ences in response and sensitivity by different species to the same toxicant were attributed to several algal characteristics including photosynthetic capacity, pigment type, cellular lipid and protein content, and cell size. Single species bioassays either for one chemical alone or in mixture have been well reported and tolerance of both marine and freshwater water-column phytoplaktonic species has been examined. Adequate published informa‐ tion on multispecies tests (communities) in laboratory and field studies exists. However, risk assessment on photosynthesis of microbenthic periphyton is inadequate, though it is essential especially for hydrophobic organic molecules. Further studies are required to evaluate the adverse effects of metabolites on aquatic microalgae.

**Keywords:** aquatic toxicology, microorganisms, chlorophyll, photosynthesis, pollutants

#### **1. Introduction**

Aquatic ecosystems receive direct or indirect inputs of a wide diversity and a variety of chemical species among which polychlorinated biphenyls (PCBs), chlorinated dioxins, polycyclic aromatic

© 2016 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited. © 2017 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

hydrocarbons (PAHs), insecticides, herbicides, oils, metals and metalloids, inorganic nonme‐ tallic elements, effluents, surfactants, synthetic detergents, and pharmaceuticals are included. Especially sediments (estuarine, river, and lake) accept the highest loads of all these aforemen‐ tioned organic and inorganic molecules in both marine and freshwater aquatic environments. As a consequence, several compounds can play the role of toxic agents that inevitably expose inhabitants of these ecosystems which are vulnerable to pollution [1].

Fortunately, over the past few decades an enormous emphasis was placed on the section of aquatic toxicological research. Environmental protection agencies in a number of countries, particularly in Europe, North America, Japan, Southeast Asia, and Australia-New Zealand, in order to deal with wastewater discharges and in addition in their efforts to curb aquatic pollution, have recognized the great value of applying aquatic hazard assessment prin‐ ciples and procedures to effluents and their component chemicals and properties.

Phototrophic microorganisms such as micro- and macroalgae contribute significantly to primary productivity, nutrient cycling, and decomposition in the aquatic ecosys‐ tems; therefore, their importance in providing energy that sustains invertebrates and fish of those environmental compartments is very crucial. Microalgal communities form an essential functional group in aquatic habitats not only as key primary producers (important food source for feeders) but as regulators of oxygen levels; even at the water sediment interface, oxygen (O<sup>2</sup> ) production is highly dependent on the photosynthesis of microphytobenthos. Thus, the effects of toxic substances on algae are important not only for those microorganisms themselves but have subsequent impacts on higher trophic lev‐ els of the food chain. Since photosynthesis forms the fundamental basis of the food webs, even sub-lethal effects on primary producers could impact the energy transfer throughout the food chain [2].

As a result, toxicity tests have been developed that assess the effects of toxicants on photosynthetic activity of exposed species. The scientific published data demonstrate that the inhibition of photosynthetic activity is a common effect parameter monitored not only in numerous laboratory toxicity tests with cultured algae but also *in situ* with natural phytoplankton and periphyton communities [3].

The focus of this chapter is to provide a review of studies describing the toxicity of various organic and inorganic contaminants on the photosynthetic apparatus of aquatic microorgan‐ isms, such as algae. It describes the biochemical mode of action of each organic and inorganic pollutant concerning the disruption of photosynthesis, discusses the methods that have been employed for its analysis, compares the sensitivities of tested algal species to various toxi‐ cants, comments on the ecological relevance of the findings, and declines areas where future research is needed to be conducted.

#### **2. Photosynthesis**

Photosynthesis is an energy transformation process that converts light energy into chemical energy and is carried out by phototrophic organisms. Photosynthesis involves a series of biochemical and biophysical reactions occurring simultaneously in photosynthetic organ‐ isms (plants, algae, and cyanobacteria) that are always starting with the absorption of pho‐ tons and ending with the incorporation of inorganic carbon into stable organic compounds called carbohydrates, such as sugars. The process of photosynthesis can be divided into two phases: the light reactions and the light independent or dark reactions. The light-dependent reactions of photosynthesis are mediated by four large protein complexes (also referred as supra- molecular complexes), embedded in the thylakoid membrane of the chloroplast: Photosystem I (PSI), Photosystem II (PSII), Cytochrome *b*<sup>6</sup> */f* Complex, and adenosine triphos‐ phate (ATP) synthase [4]. In brief, light reactions involve the excitation of electrons of chloro‐ phyll (chl) molecules within the PSII Complex to a higher energy state, which is the excited triple state (\*chl<sup>3</sup> ). This energy is harvested in the formation of several ATP molecules from ADP and inorganic phosphorus. In the PSI Complex, a similar excitation of electrons occurs, with the energy harvested to form reduced nicotinamide adenine dinucleotide phosphate (NADPH) from NADP<sup>+</sup> . The electron transfer processes involved in the light-dependent reactions of photosynthesis are depicted in **Figure 1**, which is also known as Z-scheme of photosynthesis.

hydrocarbons (PAHs), insecticides, herbicides, oils, metals and metalloids, inorganic nonme‐ tallic elements, effluents, surfactants, synthetic detergents, and pharmaceuticals are included. Especially sediments (estuarine, river, and lake) accept the highest loads of all these aforemen‐ tioned organic and inorganic molecules in both marine and freshwater aquatic environments. As a consequence, several compounds can play the role of toxic agents that inevitably expose

Fortunately, over the past few decades an enormous emphasis was placed on the section of aquatic toxicological research. Environmental protection agencies in a number of countries, particularly in Europe, North America, Japan, Southeast Asia, and Australia-New Zealand, in order to deal with wastewater discharges and in addition in their efforts to curb aquatic pollution, have recognized the great value of applying aquatic hazard assessment prin‐

Phototrophic microorganisms such as micro- and macroalgae contribute significantly to primary productivity, nutrient cycling, and decomposition in the aquatic ecosys‐ tems; therefore, their importance in providing energy that sustains invertebrates and fish of those environmental compartments is very crucial. Microalgal communities form an essential functional group in aquatic habitats not only as key primary producers (important food source for feeders) but as regulators of oxygen levels; even at the water

microphytobenthos. Thus, the effects of toxic substances on algae are important not only for those microorganisms themselves but have subsequent impacts on higher trophic lev‐ els of the food chain. Since photosynthesis forms the fundamental basis of the food webs, even sub-lethal effects on primary producers could impact the energy transfer throughout

As a result, toxicity tests have been developed that assess the effects of toxicants on photosynthetic activity of exposed species. The scientific published data demonstrate that the inhibition of photosynthetic activity is a common effect parameter monitored not only in numerous laboratory toxicity tests with cultured algae but also *in situ* with natural

The focus of this chapter is to provide a review of studies describing the toxicity of various organic and inorganic contaminants on the photosynthetic apparatus of aquatic microorgan‐ isms, such as algae. It describes the biochemical mode of action of each organic and inorganic pollutant concerning the disruption of photosynthesis, discusses the methods that have been employed for its analysis, compares the sensitivities of tested algal species to various toxi‐ cants, comments on the ecological relevance of the findings, and declines areas where future

Photosynthesis is an energy transformation process that converts light energy into chemical energy and is carried out by phototrophic organisms. Photosynthesis involves a series of

) production is highly dependent on the photosynthesis of

ciples and procedures to effluents and their component chemicals and properties.

inhabitants of these ecosystems which are vulnerable to pollution [1].

sediment interface, oxygen (O<sup>2</sup>

phytoplankton and periphyton communities [3].

research is needed to be conducted.

**2. Photosynthesis**

the food chain [2].

38 Chlorophyll

Algae during the light-dependent reactions of photosynthesis that take place in chloroplasts use pigment chl to absorb light, split the molecule of water, and therefore produce oxygen gas, and energy storage compounds of NADPH and ATP. Despite the fact that algae consti‐ tute a large, diverse, and polyphyletic group of organisms that exhibit enormous variations in morphology and physiology, the most important common biochemical attribute that unites photosynthetic algal species is their ability to perform photosynthesis.

**Figure 1.** The Z-scheme of electron transfer processes involved in the light-dependent reactions of photosynthesis.

#### **3. Methodologies of algal photosynthesis inhibition tests**

Historically since the early 1900s, a variety of toxicity tests using algal species as exposed organisms have been performed for the evaluation of phytotoxic effects of several types of potential toxicants on aquatic inhabitants (including commercial chemicals, industrial and municipal effluents, and hazardous wastes). In the early 1970s and after taking into account the enormous ecological importance of bioassays, a number of regulatory and standard development agencies such as the Organization for Economic Cooperation and Development (OECD), International Standards Organization (ISO), European Economic Community (EEC), American Public Health Association (APHA), American Society for Testing and Materials (ASTM), and US Environmental Protection Agency (USEPA) developed and standardized phytotoxicity test methods. Current test methods are designed under the assumption that effects can be studied by three general approaches: (I) in a controlled laboratory experiment with limited number of variables, (II) in an experimental model ecosystem (indoor or outdoor simulator), and finally (III) in a natural ecosystem (*in situ*) [5].

Cause and effect relationships of specific chemicals to different types of target species are easily studied by the conduction of single-species laboratory-controlled experiments. The various methodologies of single-species bioassays differ slightly in design, but basically they utilize a uni-algal population of an available, easily cultivated, and sensitive algal test species (based on these criteria several microalgae have been recommended as standard test species, such as *Selenastrum capricornutum*), which is exposed during its log-growth phase to a range of concentrations of the toxicant [6].

The main disadvantage and limitation of single-species bioassays is the fact that they focus on assessing the effects of toxicants on single species and are performed under controlled labora‐ tory conditions which are considerably different from the conditions of a realistic environment. In natural aquatic ecosystems, many complex species interactions and environmental influ‐ ences and changes that cannot be simulated in laboratory studies continually occur. Other types of laboratory-conducted toxicological studies and beyond the level of single- species test are the multispecies tests and the small ecosystem tests, which are also called laboratory microcosms, and involve small-scale enclosures that contain natural samples (water, sediment, and algae) providing a simple simulation of natural systems. Phytoplankton and periphyton are the flora utilized in most multispecies toxicity tests [7].

Natural field studies or natural aquatic ecosystems tests (pond, stream, lake, or estuary) are defined as those in which both the test system and exposure to the stressor are naturally derived [8]. Field tests are very important and reliable for evaluating and understanding the biological and ecological effects of chemicals under real environmental conditions. Outdoor microcosms or mesocosms are simulated field studies that are composed of either an isolated subsection of the natural aquatic reservoir or a man-made physical model of an aquatic ecosystem, whereas the test systems are manually treated with the test chemical at predeter‐ mined test concentrations [8]. In general, the utilization of microcosms and mesocosms for assessing the effects of toxicants can reduce the possibility of an inaccurate estimation of the adverse effects of pollutants on aquatic species belonging to different ecological categories [9].

Photosynthetic activity is considered as a significant effect parameter of a variety of toxicants on algae (physiological and morphological effects). The primary advantage of photosynthesis tests is their short duration, which is usually 2–4 h, but exposure times have also ranged from 30 min to 24 h [7]. Therefore, the inhibitory and stimulatory effects of many organic and inorganic com‐ pounds on algal photosynthesis have been determined in laboratory and field studies. According to an extended published literature, several algal biochemical parameters linked to photosynthesis process such as ATP formation, CO<sup>2</sup> fixation, O<sup>2</sup> evolution, carbon uptake (<sup>14</sup>C), and chlorophyll content have been adopted as traditional and classical indicators for the evaluation of environmen‐ tal stresses caused by many classes of various contaminants on photosynthetic algal species [10].

potential toxicants on aquatic inhabitants (including commercial chemicals, industrial and municipal effluents, and hazardous wastes). In the early 1970s and after taking into account the enormous ecological importance of bioassays, a number of regulatory and standard development agencies such as the Organization for Economic Cooperation and Development (OECD), International Standards Organization (ISO), European Economic Community (EEC), American Public Health Association (APHA), American Society for Testing and Materials (ASTM), and US Environmental Protection Agency (USEPA) developed and standardized phytotoxicity test methods. Current test methods are designed under the assumption that effects can be studied by three general approaches: (I) in a controlled laboratory experiment with limited number of variables, (II) in an experimental model ecosystem (indoor or outdoor

Cause and effect relationships of specific chemicals to different types of target species are easily studied by the conduction of single-species laboratory-controlled experiments. The various methodologies of single-species bioassays differ slightly in design, but basically they utilize a uni-algal population of an available, easily cultivated, and sensitive algal test species (based on these criteria several microalgae have been recommended as standard test species, such as *Selenastrum capricornutum*), which is exposed during its log-growth phase to a range

The main disadvantage and limitation of single-species bioassays is the fact that they focus on assessing the effects of toxicants on single species and are performed under controlled labora‐ tory conditions which are considerably different from the conditions of a realistic environment. In natural aquatic ecosystems, many complex species interactions and environmental influ‐ ences and changes that cannot be simulated in laboratory studies continually occur. Other types of laboratory-conducted toxicological studies and beyond the level of single- species test are the multispecies tests and the small ecosystem tests, which are also called laboratory microcosms, and involve small-scale enclosures that contain natural samples (water, sediment, and algae) providing a simple simulation of natural systems. Phytoplankton and periphyton are the flora

Natural field studies or natural aquatic ecosystems tests (pond, stream, lake, or estuary) are defined as those in which both the test system and exposure to the stressor are naturally derived [8]. Field tests are very important and reliable for evaluating and understanding the biological and ecological effects of chemicals under real environmental conditions. Outdoor microcosms or mesocosms are simulated field studies that are composed of either an isolated subsection of the natural aquatic reservoir or a man-made physical model of an aquatic ecosystem, whereas the test systems are manually treated with the test chemical at predeter‐ mined test concentrations [8]. In general, the utilization of microcosms and mesocosms for assessing the effects of toxicants can reduce the possibility of an inaccurate estimation of the adverse effects of pollutants on aquatic species belonging to different ecological categories [9]. Photosynthetic activity is considered as a significant effect parameter of a variety of toxicants on algae (physiological and morphological effects). The primary advantage of photosynthesis tests is their short duration, which is usually 2–4 h, but exposure times have also ranged from 30 min to 24 h [7]. Therefore, the inhibitory and stimulatory effects of many organic and inorganic com‐ pounds on algal photosynthesis have been determined in laboratory and field studies. According

simulator), and finally (III) in a natural ecosystem (*in situ*) [5].

of concentrations of the toxicant [6].

40 Chlorophyll

utilized in most multispecies toxicity tests [7].

A great progress in the area of algal photosynthesis research has been made during the last decades. Based on the fact that a proportion of the absorbed light energy in PSII photochem‐ istry cannot be used to drive electron transport and is dissipated via non-radiative energy as heat or chlorophyll fluorescence emission associated with the PSII complex [2, 11–15], informa‐ tion about changes in the efficiency of photosynthesis can be acquired by measuring the yield of Chl-α-fluorescence [2, 16]. Chl-α-fluorescence is a physical signal defined as the radiative energy evolved from de-exciting Chl-α-molecules (*λ* = 690 nm for PSII, *λ* = 740 nm for PSI) [17] that has been used as a rapid, non-intrusive, and highly sensitive bioindicator of algal stress in response to different chemicals in recent years [2, 18, 19]. Apart from their utility in determining the physiological status of photosynthesizers in the natural environment, Chl-α-fluorescencebased methods are applied in ecophysiological and toxicological studies [2]. Among the vari‐ ous fluorescence techniques, pulse amplitude modulation (PAM) fluorometry, introduced by Schreiber et al. [11], has been demonstrated as a rapid, non-invasive, reliable, economically fea‐ sible, time-saving, and accurate technique, well suited for investigating changes in photochem‐ ical efficiency of aquatic algae, that permits *in vivo* non‐destructive determination of changes in the photosynthetic apparatus much earlier than the appearance of visible damage [19]. Several types of PAM are known including the Maxi Imaging-PAM, Diving PAM, and ToxY-PAM fluo‐ rometer [2]. Numerous articles provide the efficiency of several Chl-α-fluorescence parameters that have been employed in assessing the effects of toxicants or their combinations on microal‐ gae and macroalgae (seaweeds). Detailed definitions of certain Chl-α-fluorescence parameters along with their photosynthetic importance are available in the literature [16, 20–22]. The most commonly used Chl-α-fluorescence key parameters that are becoming recognized as valid sublethal indicators of photosystem stress and have been used to examine the sub-lethal toxicity of toxicants toward a variety of microalgae are maximum quantum yield (*F*<sup>v</sup> */F*m), effective PSII quantum yield (*Φ*PSII, or *Φ*m or *ΔF/F*m'), operational PSII quantum yield (*Φ'*PSII or *Φ'*m), propor‐ tion of open PSII (*qP*), non-photochemical quenching (*NPQ*), and electron transport rate (*ETR*) [2, 23–26]. Hence, new types of devices of dual-channel PAM Chl fluorometers have been devel‐ oped, which are specialized in the detection of extremely small differences in photosynthetic activity in algae or thylakoids suspensions. In conjunction with standardized algae cultures or isolated thylakoids, they provide an ultrasensitive bioassay system occupied frequently for the detection of toxic substances in water samples [24, 27]. Furthermore, many studies have directly compared the sensitivity of Chl-α-fluorescence end points to traditional indicators of organic and inorganic chemical stress on algae; these surveys include herbicides [26], antifoul‐ ing agents, organometallic compounds [28], and metals [29].

#### **4. Oils, dispersants, and dispersed oils**

Naturally occurring raw or unprocessed crude oil and petroleum products are both included in the term "petroleum." Petroleum is a mixture of hydrocarbons of various molecular weights (most of which are alkanes, cycloalkanes, and various aromatic hydrocarbons), other organic compounds containing nitrogen, oxygen, and sulfur, and trace amounts of metals such as iron, nickel, copper, and vanadium. Hence, crude oil is a highly toxic compound comprising a mixture of up to 10,000 different types of hydrocarbons, both aliphatic and aromatic, which produce great damage to aquatic ecosystems [30]. On the other hand, processed and refined petroleum products include a large number of fuels, lubricants, and petrochemicals, such as gasoline, kerosene, diesel, paraffin wax, and many others that can cause important environ‐ mental contamination if released in ecosystems.

Hydrocarbons in aquatic environments have biogenic, natural geologic, and anthropogenic origins such as oil spills (releases of crude oil from tankers, offshore platforms, drilling rings, as well as spills of refined petroleum products and their by-products, or spills of any oil refuge or waste oil) [31–33]. Adverse effects resulting from spilled oil can be a result of (I) dissolved materials, (II) physical effects due to contact with oil droplets, (III) enhanced uptake of petro‐ leum hydrocarbons through oil/organism interactions, or (IV) a combination of these factors [34]. Besides all the above, the insoluble and mainly the soluble fractions of oil reduce light penetration into the water column affecting phytoplankton photosynthesis process [35].

The ecological effects of accidental oil spills have been the subject of relevant laboratory and field research. Since the decade of 1950s, it has been known that crude and refined oils are phytotoxic [32], whereas the scientific interest concerning the sub-lethal effects of oils and their components on enzyme systems, photosynthesis, respiration, and protein and nucleic acid synthesis of primary producers is steadily increasing nowadays. According to pub‐ lished scientific data, it is demonstrated that toxic effect concentrations for oils and algae vary greatly. As previously reported in a recent review paper, the toxic effect concentrations range is between 0.002 and 10,000 ppm for crude oils and between 0.09 and 50 ppm for refined oils [32].

Based on information presented in the same bibliographic review of Lewis et al. on toxicity of oils, dispersants (mixtures of emulsifiers and solvents that break an oil slick into smaller droplets of oil), and dispersed oils toward algae and aquatic plants, 22 species of freshwater and 63 species of saltwater algae have been exposed to more oils (21) and dispersants (27) than any other type of aquatic plant [32]. This numeric example shows that even though damage may occur from low-level continuous discharges to both freshwater and saltwater environ‐ ments, however, the environmental effects of large oil spills to marine waters have received the most attention by the public and regulatory and scientific communities resulting in the imbalance of entries in toxicity databases. Some of the available literature data concerning the toxicity of several types of oil or individual hydrocarbons on the photosynthetic apparatus reported for various algae are presented in **Table 1**.

The effects of crude oils and oil components on algae have been widely studied [43, 47–55], and among the different employed response parameters the effects on photosynthetic activity were included [43, 56, 57]. For that purpose, several algal species have been exposed to crude oils, fuel oils, dispersants, and dispersed oils not only in uni-algal cultures grown under labora‐ tory‐controlled conditions but also *in situ* as well by short- and long-term studies using micro‐ cosms, or mesocosms and mostly in short-term laboratory experiments. Toxicology studies Effects on the Photosynthetic Activity of Algae after Exposure to Various Organic and Inorganic Pollutants: Review http://dx.doi.org/10.5772/67991 43

(most of which are alkanes, cycloalkanes, and various aromatic hydrocarbons), other organic compounds containing nitrogen, oxygen, and sulfur, and trace amounts of metals such as iron, nickel, copper, and vanadium. Hence, crude oil is a highly toxic compound comprising a mixture of up to 10,000 different types of hydrocarbons, both aliphatic and aromatic, which produce great damage to aquatic ecosystems [30]. On the other hand, processed and refined petroleum products include a large number of fuels, lubricants, and petrochemicals, such as gasoline, kerosene, diesel, paraffin wax, and many others that can cause important environ‐

Hydrocarbons in aquatic environments have biogenic, natural geologic, and anthropogenic origins such as oil spills (releases of crude oil from tankers, offshore platforms, drilling rings, as well as spills of refined petroleum products and their by-products, or spills of any oil refuge or waste oil) [31–33]. Adverse effects resulting from spilled oil can be a result of (I) dissolved materials, (II) physical effects due to contact with oil droplets, (III) enhanced uptake of petro‐ leum hydrocarbons through oil/organism interactions, or (IV) a combination of these factors [34]. Besides all the above, the insoluble and mainly the soluble fractions of oil reduce light penetration into the water column affecting phytoplankton photosynthesis process [35].

The ecological effects of accidental oil spills have been the subject of relevant laboratory and field research. Since the decade of 1950s, it has been known that crude and refined oils are phytotoxic [32], whereas the scientific interest concerning the sub-lethal effects of oils and their components on enzyme systems, photosynthesis, respiration, and protein and nucleic acid synthesis of primary producers is steadily increasing nowadays. According to pub‐ lished scientific data, it is demonstrated that toxic effect concentrations for oils and algae vary greatly. As previously reported in a recent review paper, the toxic effect concentrations range is between 0.002 and 10,000 ppm for crude oils and between 0.09 and 50 ppm for refined

Based on information presented in the same bibliographic review of Lewis et al. on toxicity of oils, dispersants (mixtures of emulsifiers and solvents that break an oil slick into smaller droplets of oil), and dispersed oils toward algae and aquatic plants, 22 species of freshwater and 63 species of saltwater algae have been exposed to more oils (21) and dispersants (27) than any other type of aquatic plant [32]. This numeric example shows that even though damage may occur from low-level continuous discharges to both freshwater and saltwater environ‐ ments, however, the environmental effects of large oil spills to marine waters have received the most attention by the public and regulatory and scientific communities resulting in the imbalance of entries in toxicity databases. Some of the available literature data concerning the toxicity of several types of oil or individual hydrocarbons on the photosynthetic apparatus

The effects of crude oils and oil components on algae have been widely studied [43, 47–55], and among the different employed response parameters the effects on photosynthetic activity were included [43, 56, 57]. For that purpose, several algal species have been exposed to crude oils, fuel oils, dispersants, and dispersed oils not only in uni-algal cultures grown under labora‐ tory‐controlled conditions but also *in situ* as well by short- and long-term studies using micro‐ cosms, or mesocosms and mostly in short-term laboratory experiments. Toxicology studies

mental contamination if released in ecosystems.

reported for various algae are presented in **Table 1**.

oils [32].

42 Chlorophyll



**Table 1.** Examples of oils and hydrocarbons toxicity on the photosynthetic apparatus reported for various algae. Reports in chronological order.

conducted with photosynthetic aquatic communities usually indicate a shift of species com‐ position and abundance after an oil spill due to the replacement of sensitive species by resis‐ tant ones (observations of short-term studies) [58]. Long-term studies in most cases reported cascades of late, indirect impacts on coastal communities due to chronic exposures to environ‐ ment-sequestered petroleum products that delayed ecosystem recovery for years after an oil spill [59, 60]. Results of phytoplankton community studies are quite variable depending on characteristics of the oil, characteristics of the exposed algal species, influence of dispersants, type of ecosystem affected, dynamics of water masses, and numerous other variables [60, 61]. Therefore, the ecological impact following an oil spill depends on the volume spilled, oil type, geographical location of the spill, the characteristics of the receiving water, and its biota (e.g., sensitivity of organisms), and duration of contact with oil [62].

Short-term laboratory experiments, using laboratory-tolerant taxa and model experimental designs, have also been performed in order to evaluate more specifically the effects of dif‐ ferent petroleum products on algal photosynthesis. Toxicity data obtained from laboratory assays indicate that toxic effects depend on the phytoplanktonic species, the group of oils involved, and the physical characteristics of the water, such as concentrations of dissolved organic compounds, temperature, salinity currents, redox potential, and nutrient loading [60].

In general, responses of microscopic photosynthesizers to oil are diverse [63]. In some case studies, growth rate has been shown as a more sensitive end point parameter than photosynthetic activity [40], whereas in others Chl-α-content and carbon uptake were more sensitive parameters for assessing hydrocarbon toxicity than cell counting [39]. In our knowledge, the dominant effect observed on photosynthetic activity after exposure to petroleum hydrocarbons is inhibition, while stimulation effects at low exposure levels of the toxicants have been also reported [37, 41, 64].

These findings are in accordance with the observations that microalgae have the capability to grow in the crude oil-contaminated environments, such as in the case of the rapid adaptation of mesophile species to crude oil of the Arroyo Minero River (Argentina) [30]. Hence, microalgae are able to survive in adverse environments as a result of physiological acclimation due to the modification of gene expression [30]. However, when values of environmental stress exceed physiological limits, survival depends exclusively on adaptive evolution, which is supported by the occurrence of mutations that confer resistance [30].

#### **5. Pesticides**

**Test compounds Test species Observed stress response References**

Reduced photosynthetic capacity. Highest sensitivity: *S. costatum.* Similar results by lab batch and *in situ* dialysis culture.

Chl-α reduced at oil exposure concentration of 1–2 mg L−1; No observed affection in marine community composition.

In terms of Chl-α content: 3d EC50 = 0.015 g L−1; 5d EC50 = 0.014 g L−1;

7d EC50 = 0.0156 g L−1.

chlorophyll-α concentration decreased after 24–72 h.

photosynthesis (based on *F*<sup>v</sup>

High concentrations of oil (≥2.28 mg L− 1) of decreased

phytoplankton population, inhibition of photosynthesis associated with degradation of pigments (increase in phaeophytin).

Chl-α content.

Natural periphyton *In situ*: significant decrease in

**Table 1.** Examples of oils and hydrocarbons toxicity on the photosynthetic apparatus reported for various algae. Reports

conducted with photosynthetic aquatic communities usually indicate a shift of species com‐ position and abundance after an oil spill due to the replacement of sensitive species by resis‐ tant ones (observations of short-term studies) [58]. Long-term studies in most cases reported cascades of late, indirect impacts on coastal communities due to chronic exposures to environ‐ ment-sequestered petroleum products that delayed ecosystem recovery for years after an oil spill [59, 60]. Results of phytoplankton community studies are quite variable depending on characteristics of the oil, characteristics of the exposed algal species, influence of dispersants, type of ecosystem affected, dynamics of water masses, and numerous other variables [60, 61]. Therefore, the ecological impact following an oil spill depends on the volume spilled, oil type, geographical location of the spill, the characteristics of the receiving water, and its biota (e.g.,

*F*m, *ETR*max, and photosynthetic efficiency α-values) after only 1 h of oil exposure with clear concentration dependency. After 3 d, photosynthesis remained inhibited although cell survival was only slightly effected.

Microcosms Photosynthetic activity and

*Dunaliella tertiolecta* Significant inhibition of

Hegseth and Ostgaard

Siron et al. (1993) [43]

El-Dib et al. (1997) [44]

González et al. (2009)

Carrera-Martinez et al.

Huang et al. (2010) [45]

Jaiswar et al. (2013) [46]

[35]

(2010) [30]

*/*

(1985) [42]

*Skeletonema costatum Thalassiosira pseudonana Phaeodactylum tricornutum*

St. Laurence Estuary phytoplankton (*in situ*

dosing)

Eight groups of crude oil Marine phytoplankton

community

Note: *Pseudokirchneriella subcapitata*, known as *Selenastrum capricornutum*.

sensitivity of organisms), and duration of contact with oil [62].

*Selenastrum capricornutum*

Crude oils: Ekofisk and

Crude oil: Norman Wells

Diesel fuel oil No. 2: American Petroleum

Chrysene (water soluble

Oil samples from the tanker *Prestige* spill

Accidental oil spill in Mumbai Harbor

in chronological order.

Stratjford

44 Chlorophyll

Corexit 9550

Institute

PAH)

Pesticides are phytotoxins that are widely used all over the world in agriculture to kill unwanted vegetation. Pesticides are defined as substances or mixtures of substances intended for control‐ ling, preventing, destroying, repelling, or attracting any biological organism deemed to be a pest. Insecticides, herbicides, defoliants, desiccants, fungicides, nematicides, avicides, and rodenti‐ cides are some of the many categories of pesticides. Many members of these compounds are very selective and are applied against certain target species, whereas many others are completely non‐ selective and thus effective to almost every species of plants acting as wide-spectrum molecules.

Paradoxically, these substances do not always remain in agricultural soils where they are applied for crop protection and fruit tree treatment, but sometimes they find their way into aquatic systems through leaching, surface runoff, spray-drift, soil erosion, and volatilization. Estimates indicate that the average agricultural herbicide loss is around 1% of the applied volume [27, 65]. In addition, millions of pounds of active pesticide ingredients are applied in coastal watersheds each year and that way pesticides may affect marine inhabitants via spills, runoff, and drift [66]. As a consequence, aquatic reservoirs receive direct and indirect pesti‐ cide inputs, inevitably exposing microorganisms to pesticides.

Pesticides have been classified by scientists according to their mechanisms of action. Photosynthetic inhibitors include many chemical groups of herbicides that disrupt photosyn‐ thesis pathways by four basic mechanisms that are summarized in **Figure 2**.

**Figure 2.** Photosynthetic inhibitors and their mechanism of action.

In this point, it must be mentioned that even though the majority of the pesticides is designed to and produced in the market with the assumption that they directly affect only one pri‐ mary molecular site of action in the target organism; however, many of these compounds can cause a cascade of secondary and tertiary effects as well. For example, it has been found that most photosynthetic inhibitors also can affect plant respiration at higher doses [67]. Oxidative stress can also occur as a secondary effect of PSII inhibitors [68].

Furthermore, many non-photosynthetic inhibitors have been found to have an effect on photosynthetic process of various algal species. The herbicide flazasulfuron, a member of the chemical group of sulfonylureas, which are known to cause inhibition of amino acid synthe‐ sis, belongs to that case; bioassays conducted with the freshwater algae *Scenedesmus obliquus* revealed reduction in chlorophyll content at exposure concentration of 10 μg L−1, while the increase of pigment content was reduced with the lowest tested level of exposure (0.1 μg L−1) [69]. Moreover, studies of pesticide effects on algae showed that some pesticides can inhibit photosynthesis process with two independent mechanisms. For example, it has been reported that fluometuron, a substituted phenylurea compound, not only inhibited the production of Chl pigment in the unicellular algae *Chlorella pyrenoidosa* and *Euglena gracilis* but also blocked the biosynthesis of carotene via a process known as bleaching [70].

A broad base of toxicity data involving ecotoxicology of several classes of herbicides toward non-target microorganisms is available. Numerous reports have elaborated the impacts of various herbicides to algal photosynthetic activity. However, due to limited extent only few of them are selected to be presented herein this chapter. Therefore, only some of the available data in the literature are summarized in **Table 2** so as to depict the wide range among exposed algal species and among the employed photosynthesis parameters.

Algal species vary considerably in sensitivity to herbicides stress, and several factors may contribute to species-specific sensitivity including pigment type and photosynthetic capacity, cellular lipid and protein content, and cell size [71]. For instance, tolerance to atrazine has been linked to cell size in microalgae [71], whereas increased atrazine sensitivity to cell biovolume was observed, with smaller species being more sensitive to the herbicide [72]. What is more, algal subcellular responses to herbicides have been found to be also species dependent. In general, chlorophytes are considered to be more sensitive than bacillariophytes when comparing herbicide toxicity across phyla [73]. It has been well established that environmen‐ tal parameters (light exposure, nutrient concentrations, etc.) interfere in the responses of algal communities to pesticides [74, 75]. As reported in reference [74], diatoms were more sensitive to atrazine during light exposure, suggesting that in the context of light, the response of algae depends on the season of study and on the site where samples are taken [76]. Light history has previously been implicated in periphytic (attached) microalgae, with shade-adapted (gener‐ ally diatom-dominated) communities less susceptible than sun-adapted (chlorophyte-domi‐ nated) communities [74].

Additionally, in some species, results of algal bioassays may vary significantly based on the end point selected. As reported in a published comparative study of four estuarine microalgal species, a planktonic chlorophyte (*Dunaliella tertiolecta*), a benthic chlorophyte (*Ankistrodesmus* sp.), a cryptophyte (*Storeatula major*), and a dinoflagellate (*Amphidinium* 

**Figure 2.** Photosynthetic inhibitors and their mechanism of action.

46 Chlorophyll


Effects on the Photosynthetic Activity of Algae after Exposure to Various Organic and Inorganic Pollutants: Review http://dx.doi.org/10.5772/67991 49


Note: *Pseudokirchneriella subcapitata*, known as *Selenastrum capricornutum*.

**Pesticide (Chemical class)**

48 Chlorophyll

Glyphosate (Organophosphate)

Flazasulfuron (Sulfonylurea)

Atrazine (Triazine)

Cypermethrin (Pyrethroid)

Atrazine, simazine, hexazinone (Triazine) and diuron (Urea)

40 herbicides from 18 chemical classes and 9 modes of action

**Test species Exposure conditions, observed stress response and findings**

8.9–89 mg L−1.

*Scenedesmus obliquus* 24 or 48 h at 0.1–1000 μg L−1 (Chl-*α* and

sensitive biomarker.

*Scenedesmus obliquus* 96 h at 50–250 mg L−1 (Chl-*α* and

*Raphidocelis subcapitata* EC50 with respect to the photosynthetic

oxidase.

Short-term carbon assimilation. Exposure range: 0.89–1800 mg L−1. Photosynthetic activity decreased with increasing herbicide concentration in most ponds. Range of EC50 values:

*‐b,* carotenoids content): Reduction in chls content at 10 μg L−1, while the increase of pigment content was reduced with the lowest tested level of exposure (0.1 μg L−1). Among the three pigments studied Chl-*α* was the more

Nominal concentrations of atrazine tested: 0, 12.5, 25, 50, 100, and 200 μg L−1. Atrazine significantly decreased cell density, productivity rate, biomass, and biovolume in all the algal populations tested at atrazine concentrations ≥12.5 μg L−1. Based on photosynthetic carbon assimilation: *D. tertiolecta*: EC50 = 66.81 μg L−1; *Ankistrodesmus* sp.: EC50 = 37.07 μg L−1; *Storeatula major*: EC50 = 22.17 μg L−1; *A. operculatum:* EC50 = 33.07 μg L−1; Based on photosynthetic pigments content: *D. tertiolecta*: EC50 = 65.00 μg L−1; *Ankistrodesmus* sp.: EC50 = 11.87 μg L−1; *Storeatula major*: EC50 = 45.81 μg L−1; *A. operculatum:* EC50 = 146.71 μg L−1.

*‐b,* carotenoids content): Decreased contents of chls and carotenoids. Carotenoids production more sensitive

than the ratio of Chl-*α*/Chl*‐b*.

Based on PSII quantum yield: Atrazine: IC10 = 4.4 μL L−1; Simazine: IC10 = 29.0 μL L−1; Hexazinone: IC10 = 2.7 μL L−1; Diuron: IC10 = 0.74 μL L−1

processes ranged from 0.0007 to 4.2286 mg L−1. Descending order of the average acute toxicity was photosynthetic process>cell division>lipid synthesis, acetylcoenzyme A carboxylase>acetolactate

synthase> 5-enolpyruvylshikimate-3-phosphate-syntha-se, glutamine synthase, hormone synthesis>protoporphyrinogen

Periphytic algal communities from 6 small forest ponds

*Dunaliella tertiolecta Ankistrodesmus* sp. *Storeatula major Amphidinium operculatum*

*Phaeodactylum tricornutum*

**References**

(2003) [69]

[71]

[77]

Goldsborough et al. (1998)

Couderchet and Vernet

DeLorenzo et al. (2004)

Li et al. (2005) [79]

Bengtson Nash et al. (2005) [27]

Ma et al. (2006) [88]

**Table 2.** Examples of pesticides toxicity on the photosynthetic apparatus reported for various algae. Reports in chronological order.

*operculatum*), which were exposed to atrazine, significant differences in sensitivity were observed depending on the test end point used. Chlorophyll-α was a significantly more sen‐ sitive test end point for *Ankistrodesmus* sp., biovolume was a significantly more sensitive test end point for *A. operculatum*, and phototrophic carbon assimilation was a significantly more sensitive test end point for *S. major* and *A. operculatum* [71]. In the same survey, it is suggested that species with greater Chl-α per cell are expected to be less sensitive to PSII inhibitors, because Chl-α is directly related to the amount of PSII in the cell, which is the primary bio‐ chemical target of such insecticides, and hence the more photosynthetic targets available, the more pesticide would be required to block it [71].

A dose-dependent inhibition of photosynthetic activity of algae has been reported in cases of single species [10, 27, 71] and as well as in periphytic algae exposures to a range of insecticides concentrations [77, 78].

According to the bibliographic data, available pigments content has often been used as a classic biomarker of exposure to pesticides in plants including algae and phytoplankton [69, 79, 80]. In other cases of published ecotoxicology studies evaluating the inhibition of photosynthesis by PSII inhibitors, Chl-α-fluorescence parameters were selected instead as test end points, emphasizing the precision and time-saving virtues of the technique [10, 24, 81]. For example, the inhibition of effective quantum yield (*Φ*PSII or *ΔF/F*m') has been used by many authors in order to examine the sub-lethal toxicity of herbicides toward a variety of microalgae, with some being sensitive to diuron at environmentally relevant concentrations [24, 25, 27]. Similar sensitivities were measured using <sup>14</sup>C uptake in benthic microalgae in temperate waters [82].

Taking into account the possible interactions between substances in combination, many mix‐ ture ecotoxicological experiments were performed using binary or ternary combinations of herbicides [83, 84]. Furthermore, a large body of literature data is available concerning the prediction of the joint effect of mixtures of pesticides based on their individual impacts and specific modes of action [85, 86]. Concentration addition (CA) and independent action (IA) model are the most commonly used models to predict mixture effects for similar- and dissimi‐ lar-acting compounds, respectively. Both theories assume enhanced effects with an increasing number of compounds and non-interaction between substances. Therefore, a deviation from the prediction indicates antagonism (weaker effects than predicted) or synergism (stronger effects) [87].

Pesticides are probably the most well-studied chemical group within ecotoxicological mixtures studies. This is not only due to the use of chemical mixtures in pesticide formu‐ lations and tank mixtures and the resulting co-occurrence in agricultural areas, but just as much because of the in-depth knowledge of their physiological mode of action [87]. These facts make them ideal candidates for testing mixture models based on the chemical mode of action and understanding the physiological mechanisms behind possible interactions [85, 90]. Mixture toxicity studies focused on single species [85, 86], natural communities in laboratory experiments [3, 82, 91], or outdoor microcosms and mesocosms [83, 92–94] data. Many reviews and critical analysis have shown that synergistic interactions within pesticide mixtures and realistic low-dose chemical mixtures in species are a rather rare phenomenon, constituting very low percentages of the tested mixture combinations and often occurs at high concentrations [87, 95–101]. According to the results of a comprehensive systematic review in which cocktail effects and synergistic interactions of chemicals in mixtures were predicted, synergy phenomena occurred only in 7% of the 194 binary pesticide mixtures included in the data compilation on frequency [101] (the database of Belden et al. [98] provided data on 207 pesticide mixtures of which 194 were binary and another 13 consisted of more than two pesticides). Results of the same study showed that PSII herbicides did not induce synergy in any of the 33 mixtures performed on algae in the pesticide database [101].

#### **6. Antifouling biocides**

Antifouling biocides are chemical substances that deter the microorganisms responsible for biofouling. Biofouling or biological fouling is the accumulation of microorganisms, plants, algae, or animals on wetted surfaces; hence, it can occur almost anywhere where water is present (marine vessels, swimming pools, drinking water and liquid lines for cooling elec‐ tronics, medical devices and membranes, etc.). Biofouling takes place on surfaces after the formation of a biofilm that creates a surface onto which successively larger microorganisms can attach. Specifically designed antifouling materials and coatings/paints have the ability to remove or prevent biofouling by any number of organisms on such surfaces.

Antifouling biocides are introduced to antifouling paints in order to improve their efficacy against photosynthetic organisms [2]. The biocides often target the microorganisms which create the ini‐ tial biofilm, typically bacteria. Other biocides are toxic to larger organisms in biofouling, such as the fungi and algae. Many different booster biocides have been currently added to antifouling paints including tributyltin (TBT), 2-methylthio-4-tetr-butylamino-6- cyclopropylamino-s-triazine (Irgarol 1051), 4,5-dichloro-2-n-octyl-4-isothiazolin-3-one (Sea-Nine 211), diuron, cuprous oxide, chlorothalonil, zinc pyrithione, dichlofluanid, 2,3,3,6-tetrachloro-4methylsulfonyl (TCMS), pyri‐ dine, 2-(thiocyanomethylthio) benzothiazole (TCMTB), and zineb [102].

order to examine the sub-lethal toxicity of herbicides toward a variety of microalgae, with some being sensitive to diuron at environmentally relevant concentrations [24, 25, 27]. Similar sensitivities were measured using <sup>14</sup>C uptake in benthic microalgae in temperate waters [82]. Taking into account the possible interactions between substances in combination, many mix‐ ture ecotoxicological experiments were performed using binary or ternary combinations of herbicides [83, 84]. Furthermore, a large body of literature data is available concerning the prediction of the joint effect of mixtures of pesticides based on their individual impacts and specific modes of action [85, 86]. Concentration addition (CA) and independent action (IA) model are the most commonly used models to predict mixture effects for similar- and dissimi‐ lar-acting compounds, respectively. Both theories assume enhanced effects with an increasing number of compounds and non-interaction between substances. Therefore, a deviation from the prediction indicates antagonism (weaker effects than predicted) or synergism (stronger

Pesticides are probably the most well-studied chemical group within ecotoxicological mixtures studies. This is not only due to the use of chemical mixtures in pesticide formu‐ lations and tank mixtures and the resulting co-occurrence in agricultural areas, but just as much because of the in-depth knowledge of their physiological mode of action [87]. These facts make them ideal candidates for testing mixture models based on the chemical mode of action and understanding the physiological mechanisms behind possible interactions [85, 90]. Mixture toxicity studies focused on single species [85, 86], natural communities in laboratory experiments [3, 82, 91], or outdoor microcosms and mesocosms [83, 92–94] data. Many reviews and critical analysis have shown that synergistic interactions within pesticide mixtures and realistic low-dose chemical mixtures in species are a rather rare phenomenon, constituting very low percentages of the tested mixture combinations and often occurs at high concentrations [87, 95–101]. According to the results of a comprehensive systematic review in which cocktail effects and synergistic interactions of chemicals in mixtures were predicted, synergy phenomena occurred only in 7% of the 194 binary pesticide mixtures included in the data compilation on frequency [101] (the database of Belden et al. [98] provided data on 207 pesticide mixtures of which 194 were binary and another 13 consisted of more than two pesticides). Results of the same study showed that PSII herbicides did not induce synergy

in any of the 33 mixtures performed on algae in the pesticide database [101].

remove or prevent biofouling by any number of organisms on such surfaces.

Antifouling biocides are chemical substances that deter the microorganisms responsible for biofouling. Biofouling or biological fouling is the accumulation of microorganisms, plants, algae, or animals on wetted surfaces; hence, it can occur almost anywhere where water is present (marine vessels, swimming pools, drinking water and liquid lines for cooling elec‐ tronics, medical devices and membranes, etc.). Biofouling takes place on surfaces after the formation of a biofilm that creates a surface onto which successively larger microorganisms can attach. Specifically designed antifouling materials and coatings/paints have the ability to

effects) [87].

50 Chlorophyll

**6. Antifouling biocides**

One of the most commonly used biocides, and anti-fouling agents, is TBT. It is toxic to both microorganisms and larger aquatic organisms [103]. The mechanism of action of the TBT in algae is based on its interference with energy metabolism in chloroplasts and mitochondria, but it is also shown that TBT interacts with proteins and membranes and binds to or interacts with any protein containing free sulfhydryl groups [3, 104]. Bioassays conducted with the marine algae *Tetraselmis suecica* revealed that in chronic exposure to TBT, at higher concentra‐ tions (0.5–1 μg mL−1) growth rate, chlorophyll pigments, carbohydrate, and protein contents were reduced [105]. Different responses have been described among three species of marine microalga *T. tetrathele*, *Nannochloropsis oculata,* and *Dunaliella* sp., which were exposed to three concentrations of TBT (0.1, 0.5, and 1μg L−1). For *T. tetrathele*, exposure to TBT resulted in an increase of chlorophyll contents, even up to 210 and 225% at highest concentration of TBT (1μg L−1) for chlorophyll α and b, respectively. However, acquired results for the other two algal species, *N. oculata* and *Dunaliella* sp., showed that stimulation effects occurred only at the lowest concentration tested (0.1 μg L−1), as chlorophyll contents decreased at higher exposure levels, whereas *N. oculata* was the most sensitive microalga [106]. Similar results had been published in a previous study by Sidharthan et al. in which photosynthetic pigment content of the marine eustigmatophyte *N. oculata* was significantly affected, especially at elevated TBT concentrations. The same authors found that Chl-α content decreased more than 50% at TBT concentrations above 0.50 nM level, whereas at high concentration of 4 nM both the pigments were completely leached. Comparatively, carotenoid content was less inhibited by TBT toxicity (*r* = 0:917; *P* < 0.05) [107]. Reduction (60%) in the net photosynthetic activity of *Ruppia maritina* (seagrass) in TBT-spiked and impacted sediments was measured [108]. In a microcosm approach survey that was designed to study the combined effects of TBT from antifouling paints and ultraviolet-B radiation (UVΒR: 280–320 nm), on a natural planktonic assemblage (<150 μm) isolated from the St. Lawrence Estuary (eastern Canada), it was demonstrated that phytoplankton cells were affected in their physiological functions, such as their photosynthetic efficiency. According to the obtained experimental data, the reduction in the maximum quantum yield (*F*<sup>v</sup> */F*m) values were due to damage of PSII reaction centers and inhibition of ATP synthesis. Moreover, results clearly showed that the combination of TBT and UVBR stresses has synergistic effects affecting the first trophic levels of the marine food web [28]. Finally, the inhibition of photosynthesis of periphyton community has been observed after exposure to TBT (EC50 = 0.02 mg L−1) [3].

Irgarol 1051 is a triazine herbicide that has been described as an inhibitor of algal photosyn‐ thesis. More specifically, it belongs in PSII inhibitors, as it results in oxidative stress, including photooxidation of chlorophyll [109], and inhibition of the photosynthetic electron transport in chloroplasts by binding to the D<sup>1</sup> protein [110]. Irgarol 1051 was introduced after the restric‐ tions on using TBT in antifouling paints (as a replacement) [111] and has found its application as an algicide in antifouling paints for boats and vessels. Irgarol is the most hydrophobic compound of the family of the triazines due to the presence of both tert-butyl group and the cyclopropyl group [102]. It is mainly used in combination with copper [3] and is the most fre‐ quently detected antifouling biocide worldwide [102]. Even though Irgarol 1051 is a relatively new compound, several papers have been published in the last years dealing with its ecotoxi‐ cological behavior toward non-target microorganisms. For example, in algal symbionts iso‐ lated from *M. mirabilis, D. strigosa*, and *F. fragum* 40–50% reduction of net <sup>14</sup>C incorporation has been demonstrated after their 6-h exposure to 10 mg L−1 of Irgarol 1051 [112]. Inhibition of the algal photosynthetic activity of several algal species including *D. tertiolecta, Synechococcus* sp., *E. huxleyi, Fucus vesiculosus, Enteromorpha intestinalis*, *Ulva intestinalis,* and seagrass *Z. marina* by Irgarol 1051 has been summarized [113]. In addition, the destruction of periphyton photosynthesis process after exposure to the same biocide has been demonstrated (EC50 = 0.82 nM) [114]. According to the available data, Irgarol 1051 has the potential to affect the *F*<sup>ν</sup> */F*m of phytoplankton even at very low (0.03 μg L−1) environmentally relevant concentrations [115]. This conclusion is in accordance with the assumption that Irgarol 1051 concentration up to 0.23 mg L−1 negatively impacted the photosynthetic activity of the green alga *U. intestinalis* [116]. The effect of Irgarol on the values of several Chl-α-fluorescences parameters for numer‐ ous freshwater and marine algal species has been reported including the following data: according to *F*<sup>v</sup> */F*m values: EC50 = 0.33 mg L−1 for *T. weissflogii*; EC50 = 0.60 mg L−1 for *E. huxleyi*; EC50 = 0.23 mg L−1 for *Tetraselmis* sp.; EC50 = 0.11 mg L−1 for *F. japonica* [117], reduction of *F*<sup>ν</sup> */F*<sup>m</sup> values in the presence of high concentrations for *Potamogeton pectinatus* [118]; whereas accord‐ ing to *Φ*PSII or *ΔF/F*m' values: 72 h EC50 = 0.327 mg L−1 for *T. weissflogii*; 72 h EC50 = 0.604 mg L−1 for *Emiliania huxleyi*; 72 h EC50 = 0.230 mg L−1 for *Tetraselmis* sp.; 72 h EC50 = 0.110 mg L−1 for *Fibrocapsa japonica* [119]; 72 h EC50 = 0.17 mg L−1 for *H. banksii* [120]; and 72 h EC50 = 2500 ng L−1 for *E. intestinalis* [121].

The other most commonly detected biocide in areas of high boating activity is diuron (phenylurea herbicide) [102]. The toxic effects of diuron on the photosynthetic apparatus of different algal species have been examined by many authors [10, 24–27, 89, 93, 115, 117] and among other ecotoxicological data the values of IC10 = 0.74 μL L−1 (based on PSII quan‐ tum yield) for *Phaeodactylum tricornutum* [27] and IC50 = 7 μL L−1 (based on *Φ*m, 1.5 h) for *S. capricornutum* [10] are included. Natural periphyton studies have reported an induced increase in Chl-α content after long-term (29 days) exposure to low concentrations (1μg L−1) of diuron [122]. This observation is in agreement with other previous studies of Tlili et al., who found that periphyton chronically exposed to 1 μg L−1 of diuron showed higher Chl-α pigments and carbon incorporation rates than control periphyton from day 21 to day 32 of their microcosm experiment [123]. That was confirmed in a more recent survey conducted in two series of two lotic outdoor mesocosms exposed to mixture of diuron and tebuconazole (triazole fungicide) which revealed induced tolerance to diuron, and therefore it was indicated that the effects of pulsed acute exposures to pesticides on periphyton depended on whether the communities had previously been exposed to the same stressors or not [89].

It has become well known that the antifouling biocide Sea-Nine 211 has an impact as an inhibitor of PSII electron transport [2, 113]. In addition, like other, more water-soluble rep‐ resentatives from the so-called Kathon group of biocides, Sea-Nine 211 quickly penetrates cell membranes and inhibits specific enzymes in the cell by reacting with intracellular thiols [3, 124]. Sea-Nine also seems to be able to affect more than one thiol group by generating a cascade of intracellular radicals [3]. Based on *F*<sup>v</sup> */F*m measurements of natural phytoplankton communities, the toxicity of few biocides has been ranked as follows: Irgarol 1051 > zinc pyrithione>Sea-Nine 211>diuron. Thereby, it is suggested that Sea-Nine is more toxic than diuron, but less toxic than Irgarol [115]. In another survey, the toxicity of the antifoulants Sea-Nine, Irgarol, and TBT has been determined individually and in mixtures in two tests with microalgae and the effects on periphyton community photosynthesis and reproduction of the unicellular green algae *S. vacuolatus* have been investigated. The tested antifoulants have been found to be highly toxic in both tests. Observed mixture toxicities were compared with predictions derived from two concepts: independent action (IA) and concentration addition (CA), and IA failed to provide accurate predictions of the observed mixture toxicities. Mixture effects at high concentrations were slightly overestimated and effects at low concentrations were slightly underestimated [3].

Synergistic interactions have been foreseen not only between irgarol and diuron but between irgarol and chlorothalonil or 2-(thiocyanomethylthio)benzothiazole (TCMTB) as well. The synergies between irgarol and the two general fungicides, chlorothalonil and TCMTB, could be similar to the mechanism proposed for the PSII/metal interactions, as both fungicides cre‐ ate reactive oxygen species (ROS) and additionally chlorothalonil conjugates with glutathi‐ one, an important ROS scavenger [101].

#### **7. Heavy metals and metalloids**

chloroplasts by binding to the D<sup>1</sup>

52 Chlorophyll

according to *F*<sup>v</sup>

for *E. intestinalis* [121].

same stressors or not [89].

protein [110]. Irgarol 1051 was introduced after the restric‐

*/F*m of

*/F*<sup>m</sup>

tions on using TBT in antifouling paints (as a replacement) [111] and has found its application as an algicide in antifouling paints for boats and vessels. Irgarol is the most hydrophobic compound of the family of the triazines due to the presence of both tert-butyl group and the cyclopropyl group [102]. It is mainly used in combination with copper [3] and is the most fre‐ quently detected antifouling biocide worldwide [102]. Even though Irgarol 1051 is a relatively new compound, several papers have been published in the last years dealing with its ecotoxi‐ cological behavior toward non-target microorganisms. For example, in algal symbionts iso‐ lated from *M. mirabilis, D. strigosa*, and *F. fragum* 40–50% reduction of net <sup>14</sup>C incorporation has been demonstrated after their 6-h exposure to 10 mg L−1 of Irgarol 1051 [112]. Inhibition of the algal photosynthetic activity of several algal species including *D. tertiolecta, Synechococcus* sp., *E. huxleyi, Fucus vesiculosus, Enteromorpha intestinalis*, *Ulva intestinalis,* and seagrass *Z. marina* by Irgarol 1051 has been summarized [113]. In addition, the destruction of periphyton photosynthesis process after exposure to the same biocide has been demonstrated (EC50 = 0.82 nM) [114]. According to the available data, Irgarol 1051 has the potential to affect the *F*<sup>ν</sup>

phytoplankton even at very low (0.03 μg L−1) environmentally relevant concentrations [115]. This conclusion is in accordance with the assumption that Irgarol 1051 concentration up to 0.23 mg L−1 negatively impacted the photosynthetic activity of the green alga *U. intestinalis* [116]. The effect of Irgarol on the values of several Chl-α-fluorescences parameters for numer‐ ous freshwater and marine algal species has been reported including the following data:

EC50 = 0.23 mg L−1 for *Tetraselmis* sp.; EC50 = 0.11 mg L−1 for *F. japonica* [117], reduction of *F*<sup>ν</sup>

values in the presence of high concentrations for *Potamogeton pectinatus* [118]; whereas accord‐ ing to *Φ*PSII or *ΔF/F*m' values: 72 h EC50 = 0.327 mg L−1 for *T. weissflogii*; 72 h EC50 = 0.604 mg L−1 for *Emiliania huxleyi*; 72 h EC50 = 0.230 mg L−1 for *Tetraselmis* sp.; 72 h EC50 = 0.110 mg L−1 for *Fibrocapsa japonica* [119]; 72 h EC50 = 0.17 mg L−1 for *H. banksii* [120]; and 72 h EC50 = 2500 ng L−1

The other most commonly detected biocide in areas of high boating activity is diuron (phenylurea herbicide) [102]. The toxic effects of diuron on the photosynthetic apparatus of different algal species have been examined by many authors [10, 24–27, 89, 93, 115, 117] and among other ecotoxicological data the values of IC10 = 0.74 μL L−1 (based on PSII quan‐ tum yield) for *Phaeodactylum tricornutum* [27] and IC50 = 7 μL L−1 (based on *Φ*m, 1.5 h) for *S. capricornutum* [10] are included. Natural periphyton studies have reported an induced increase in Chl-α content after long-term (29 days) exposure to low concentrations (1μg L−1) of diuron [122]. This observation is in agreement with other previous studies of Tlili et al., who found that periphyton chronically exposed to 1 μg L−1 of diuron showed higher Chl-α pigments and carbon incorporation rates than control periphyton from day 21 to day 32 of their microcosm experiment [123]. That was confirmed in a more recent survey conducted in two series of two lotic outdoor mesocosms exposed to mixture of diuron and tebuconazole (triazole fungicide) which revealed induced tolerance to diuron, and therefore it was indicated that the effects of pulsed acute exposures to pesticides on periphyton depended on whether the communities had previously been exposed to the

*/F*m values: EC50 = 0.33 mg L−1 for *T. weissflogii*; EC50 = 0.60 mg L−1 for *E. huxleyi*;

In general, heavy metals are defined as metals with relatively high densities, atomic weights, or atomic numbers. On the basis of density, the term "heavy metal" is used for the elements that possess a density value greater than 4.5–5 g cm−3, such as silver (Ag), arsenic (As), cadmium (Cd), copper (Cu), mercury (Hg), manganese (Mn), nickel (Ni), lead (Pb), and zinc (Zn), while metalloid is the definition of a chemical element that has properties intermedi‐ ate between metals and non-metals, such as germanium(Ge), antimony (Sb), selenium (Se), tellurium (Te), polonium (Po), technetium (Tc), and astatine (At) [125].

Several metals are essential for living beings at very low concentrations, but at higher doses most of them are toxic for organisms belonging to different levels of the food chain [126]. Based on that criterion, metals are separated into the three following classes:

• **The essentials (class A)**: calcium (Ca), magnesium (Mg), Mn, potassium (K), sodium (Na), and strontium (Sr) (including macroelements which are metals that are required for algal growth, metabolism, and physiology (e.g., K and Mg) and microelements, which are metals that are required in trace amounts for certain biological processes and therefore must be obtained from the external environment).


With regard to Ecotoxicology and Environmental Science, the term "heavy metals" is used to refer to metals that have caused environmental problems and includes chemical elements from the non‐essentials and the borderline classes.

A steadily growing interest in the investigations on heavy metals is recorded and a large number of scientific surveys focused on the speciation of metals, their toxicity, accumulation, biomagnification, bioindication, migration, removal, phytoremediation, and biomonitoring have been conducted during the last decades. Cd, Hg, Zn, Cu, Ni, Cr, Pb, Co, V, titanium (Ti), Fe, Mn, Ag, and Sn are the metals that have been studied more extensively, whereas Hg, Cd, and Pb are some of the elements that have received the most scientific attention, possibly due to their highly toxic properties and their effectiveness on the environment and the living organisms [128].

Heavy metals can be naturally produced in aquatic system by the slow leaching from soil to water, usually at low levels [129]. Several other large natural inputs of heavy metals into water ecosystems are from the erosion or rocks, wind-blowing dusts, volcanic activity, and forest fires [128]. In addition, several anthropogenic activities such as energy production technologies, industrial effluents, and wastes (from coal mines, thermal power plants, metal‐ lurgy, plating, chemical plant, curry and paper-making industries, and other allied indus‐ tries) alter the physicochemical characteristics of water bodies and elevate the heavy metals concentration according to the nature of effluent being discharged [130, 131]. Therefore, aquatic ecosystems receive inputs of different source containing a variety of metal ions (Mx+) that are directly or indirectly discharged into them.

Aquatic plants assimilate easily heavy metals, which are strongly phytotoxic and pose a threat to freshwater and marine life. Moreover, it has been well established that, depending on its bioaccumulation characteristics, a heavy metal can disperse through the various trophic levels of an ecosystem and its concentration levels are magnified [129]. Metals are not acces‐ sible to plants in their elemental forms (valence state of 0). On the contrary, they are available only in solution; hence, only metal ions play a role in biological systems [132]. The toxic‐ ity of metals and their compounds, however, largely depends on their bioavailability, that is, the mechanisms of uptake through cell membranes, intracellular distribution, and bind‐ ing to cellular macromolecules [133]. In other words, the bioavailability of the metal, which depends on both biological factors and on the physicochemical properties of metallic forms (elements, their ions, and their compounds), is one of the key parameters in the assessment of the potential toxicity of metallic elements and their compounds toward organisms [125]. Metal availability is strongly dependent on environmental components, such as pH, redox and organic content, and soluble and bio-available metals. Hence, metals in the environment can be divided into two classes: (I) bio-available (soluble, non-sorbed, and mobile) and (II) non-bio-available (precipitated, complexed, sorbed, and no mobile).

Heavy metals enter algal cells by means of either active transport or endocytosis through chelating proteins and affect various physiological and biochemical processes of the algae. The mechanisms by which metals exert their toxicity on algae are very diverse and depend on the algal species, the nature and concentration of the metal, and the environmental con‐ ditions accompanying heavy metal stress [134]. Generally, their toxicity toward algal cells primarily results from (I) direct binding to the sulfhydryl groups (−SH) in functional proteins which disrupts their structure and function, and thus renders them inactive; (II) displace‐ ment of essential cations from specific binding sites that lead to a collapse of function; and (III) generation of reactive oxygen species, which consequently damages the macromole‐ cules [126, 135].

• **The non‐essentials (class B)**: Cd, Cu, Hg, and Ag.

from the non‐essentials and the borderline classes.

that are directly or indirectly discharged into them.

non-bio-available (precipitated, complexed, sorbed, and no mobile).

and tin (Sn) [127].

54 Chlorophyll

organisms [128].

• **The borderline class**: Zn, Pb, iron (Fe), chromium (Cr), cobalt (Co), Ni, As, vanadium (V),

With regard to Ecotoxicology and Environmental Science, the term "heavy metals" is used to refer to metals that have caused environmental problems and includes chemical elements

A steadily growing interest in the investigations on heavy metals is recorded and a large number of scientific surveys focused on the speciation of metals, their toxicity, accumulation, biomagnification, bioindication, migration, removal, phytoremediation, and biomonitoring have been conducted during the last decades. Cd, Hg, Zn, Cu, Ni, Cr, Pb, Co, V, titanium (Ti), Fe, Mn, Ag, and Sn are the metals that have been studied more extensively, whereas Hg, Cd, and Pb are some of the elements that have received the most scientific attention, possibly due to their highly toxic properties and their effectiveness on the environment and the living

Heavy metals can be naturally produced in aquatic system by the slow leaching from soil to water, usually at low levels [129]. Several other large natural inputs of heavy metals into water ecosystems are from the erosion or rocks, wind-blowing dusts, volcanic activity, and forest fires [128]. In addition, several anthropogenic activities such as energy production technologies, industrial effluents, and wastes (from coal mines, thermal power plants, metal‐ lurgy, plating, chemical plant, curry and paper-making industries, and other allied indus‐ tries) alter the physicochemical characteristics of water bodies and elevate the heavy metals concentration according to the nature of effluent being discharged [130, 131]. Therefore, aquatic ecosystems receive inputs of different source containing a variety of metal ions (Mx+)

Aquatic plants assimilate easily heavy metals, which are strongly phytotoxic and pose a threat to freshwater and marine life. Moreover, it has been well established that, depending on its bioaccumulation characteristics, a heavy metal can disperse through the various trophic levels of an ecosystem and its concentration levels are magnified [129]. Metals are not acces‐ sible to plants in their elemental forms (valence state of 0). On the contrary, they are available only in solution; hence, only metal ions play a role in biological systems [132]. The toxic‐ ity of metals and their compounds, however, largely depends on their bioavailability, that is, the mechanisms of uptake through cell membranes, intracellular distribution, and bind‐ ing to cellular macromolecules [133]. In other words, the bioavailability of the metal, which depends on both biological factors and on the physicochemical properties of metallic forms (elements, their ions, and their compounds), is one of the key parameters in the assessment of the potential toxicity of metallic elements and their compounds toward organisms [125]. Metal availability is strongly dependent on environmental components, such as pH, redox and organic content, and soluble and bio-available metals. Hence, metals in the environment can be divided into two classes: (I) bio-available (soluble, non-sorbed, and mobile) and (II) At the sub-lethal level, heavy metals can interact with the vital process of photosynthesis. Interference of heavy metals with the photosynthesis of algae is a subject of intensive research that has been well documented. Almost all heavy metals are known to cause a negative impact on nearly all the components of the photosynthetic apparatus of primary producers [2, 132]. Direct effects of heavy metals on light and dark reactions and indirect effects resulting in the decrease of the photosynthetic pigment (including chlorophyll and carotenoid) con‐ tent, as well as changes in stomata function, have been reported in the literature [132, 136]. Additionally, ions of heavy metal can damage the chloroplast membrane structure, disturb the light-harvesting and oxygen-evolving complexes, inhibit the photosystems and constitu‐ ents of the photosynthetic electron transport chain, and also block the reductive pentose phos‐ phate cycle [132, 137]. Moreover, toxic metals cause the inhibition of enzyme activities that are important in photosynthetic pathway. For example, it was found that Cd2+, Zn2+, and Hg2+ inhibited the NADP-oxidoreductase in *Euglena*, thereby significantly lowering the cell supply of NADPH [138], whereas Cu2+ was shown to inhibit plasma membrane H<sup>+</sup> -ATPase activity in *Nitella flexis* [139]. Several enzymes involved in the Calvin cycle are also inhibited, especially Rubisco (bisphosphate carboxylase oxygenase) and PEPcarboxylase [132, 136]. Reaction of heavy metals with the enzyme-SH groups in proteins, substitution of essential ions, enhance‐ ment of photoinhibition and oxidative stress, impediment of plastocyanin function, change in lipid metabolisms, and disturbances in the uptake of essential microelements are other phenomena revealed due to heavy metal exposure [140, 141]. For instance, Cu2+ and Zn2+ substituted the Mg2+ in Chl molecules bound predominantly in the light-harvesting complex II of Chlorophyta, thereby impeding the PSII reaction centers, such as in the green alga *S. quadricauda* [141].

Finally, many heavy metals have been reported to influence the photosynthetic activity of algae through bleaching process. The observed bleaching effects have been connected with the tendency of toxic metals to generate ROS, such as singlet oxygen (<sup>1</sup> O2 ) and the hydroxyl radical (\*OH), which can attack thylakoid lipids and initiate oxidation biochemical reactions that destroy membranes and damage structural pigment-protein complexes. For example, the toxicity of Cr6+ compounds has been traced to the reactive intermediates (formation of \*OH radicals from H<sup>2</sup> O2 via a Fenton reaction) generated during the reduction of Cr by living cells [142]. As observed in the case of *Chlamydomonas reinhardtii* [134], this toxic metal tends to generate ROS, which can attack thylakoid lipids (mainly unsaturated fatty acids). This initiates peroxyl-radical chain reactions, destroying membranes and damaging indirectly structural pigment-protein complexes located in chloroplast membranes [2].

According to the numerous reported data on the photosynthesis inhibition by metals, three main experimental approaches can be distinguished: (I) results obtained from experiments with isolated chloroplasts or enzymes, to which heavy metals were supplied in the assay medium, (II) data acquired from experiments performed on excised leaves, exposed to a solu‐ tion of the heavy metal, and (III) comparative laboratory experiments conducted on intact higher plants or algae, grown in a control medium and on a substrate enriched with heavy metals [140]. A summary of selected references on the toxicity of metals toward the photosyn‐ thetic apparatus for various microalgae is presented in **Table 3**.

Mercury is considered as the most toxic element among those having"no known physiological function" in algae. Based on results of ecotoxicological studies, Hg is recognized globally as an important pollutant and a serious threat to ecosystems. Hg and its compounds are persistent, bioaccumulative, and toxic. Inorganic Hg is the most common form of Hg released in the aquatic environment by industries [133]. Organic forms of Hg, such as methylmer‐ cury, revealed to have much stronger inhibitory effect than the inorganic mercury chloride on photosynthetic process [143]. Hg is able to alter the photosynthetic machinery including the chloroplastic PSI reaction center, subunit PSII, the oxygen-evolving protein, and the chloroplastic ATP synthase β-subunit [133, 144]. High levels of Hg in the form of Hg2+ have strong phytotoxic effects and when present in toxic concentrations can induce visible injuries and physiological disorders in plant cells triggering the production of ROS leading to cellular disruption [133].


Effects on the Photosynthetic Activity of Algae after Exposure to Various Organic and Inorganic Pollutants: Review http://dx.doi.org/10.5772/67991 57

initiates peroxyl-radical chain reactions, destroying membranes and damaging indirectly

According to the numerous reported data on the photosynthesis inhibition by metals, three main experimental approaches can be distinguished: (I) results obtained from experiments with isolated chloroplasts or enzymes, to which heavy metals were supplied in the assay medium, (II) data acquired from experiments performed on excised leaves, exposed to a solu‐ tion of the heavy metal, and (III) comparative laboratory experiments conducted on intact higher plants or algae, grown in a control medium and on a substrate enriched with heavy metals [140]. A summary of selected references on the toxicity of metals toward the photosyn‐

Mercury is considered as the most toxic element among those having"no known physiological function" in algae. Based on results of ecotoxicological studies, Hg is recognized globally as an important pollutant and a serious threat to ecosystems. Hg and its compounds are persistent, bioaccumulative, and toxic. Inorganic Hg is the most common form of Hg released in the aquatic environment by industries [133]. Organic forms of Hg, such as methylmer‐ cury, revealed to have much stronger inhibitory effect than the inorganic mercury chloride on photosynthetic process [143]. Hg is able to alter the photosynthetic machinery including the chloroplastic PSI reaction center, subunit PSII, the oxygen-evolving protein, and the chloroplastic ATP synthase β-subunit [133, 144]. High levels of Hg in the form of Hg2+ have strong phytotoxic effects and when present in toxic concentrations can induce visible injuries and physiological disorders in plant cells triggering the production of ROS leading to cellular

structural pigment-protein complexes located in chloroplast membranes [2].

**Metallic form Test species Observed stress response References**

*Scenedesmus obliquus* Inhibition of PSII

*F*o

*perminuta*.

Under low irradiance heavy metal substitution of Mg in chl molecules bound predominantly in PSII of Chlorophyta; Under high irradiance the chls were inaccessible to substitution and the damage occurred in the PSII reaction center instead.

photochemistry. Among the fluorescence parameters measured (after 12 h: *F*<sup>o</sup>

*qN*, *qP* and after 1 h: *F*m, *F*<sup>v</sup>

increased at low levels and inhibited in high levels. Photosynthetic electron transport in *M. minutum* was more sensitive to Co2+ than in *N.* 

Pigment content and photosynthetic O<sup>2</sup>

*/F*m) the highest sensitivity to all the five test metals had *F*<sup>v</sup>

, *F*<sup>v</sup> */F*m,

evolution:

*/2*, and

[29]

*/F*m.

Kupper et al. (2002) [141]

Mallick and Mohn (2003)

El-Sheekh et al. (2003) [156]

thetic apparatus for various microalgae is presented in **Table 3**.

disruption [133].

56 Chlorophyll

Cu2+, Ni2+, Cd2+, Zn2+,

Cr6+

Cu2+, Zn2+ *Scenedesmus quadricauda*

Co2+ *Monoraphidium minutum*

*Nitzschia perminuta*

*Antithamnion plumula Ectocarpus siliculosus*



Note: *Pseudokirchneriella subcapitata*, known as *Selenastrum capricornutum*.

**Table 3.** Examples of metals toxicity on the photosynthetic apparatus reported for various algae. Reports in chronological order.

Copper is unquestionably an essential element in various metabolic processes of algae, such as amine oxidase and cytochrome c oxidase system, prosthetic group of the chloroplastic antioxidant enzyme Cu/Zn superoxide dismutase, and regulator of PSII-mediated electron transport. However, Cu is still considered as one of the most toxic heavy metal ions to algae and is a potent inhibitor of photosynthesis [2]. Many studies have examined ecotoxicological effects of Cu on photosynthetic activity of plants and phytoplankton [145]. From an evaluation of the literature, Cu can affect photosynthetic electron transport on the reducing side of PSI at the level of the ferredoxin [146], alter the PSII on the oxidizing side by inhibiting the electron transport at P680 (the primary donor of PSII) or by inactivating some PSII reaction centers [147]. Cu may also impair the PSII electron transport on its reducing side by affecting the rate of oxidoreduction [148]. The inhibitory effect of copper on the photosynthetic apparatus of several species of algae has been examined, including *E. gracilis* [149, 150], *S. quadricauda* [141], *S. obliquus* [151], *S. incrassatulus* [152], *C. pyrenoidosa* [153], *C. vulgaris* [154], *Planothidium lan‐ ceolatum* and *Isochrysis galbana* [155], *D. tertiolecta, Promocentrum minimum, Synechococcus* sp*.,* and *Thalassiosira weissflogii* [145].

Cadmium is a heavy metal that occurs naturally in ores along with zinc, lead, and copper. Its compounds are used as stabilizers in PVC products, color pigment, several alloys, and in rechargeable nickel-cadmium batteries. Cd forms complexes with various organic particles and thereby triggers a wide range of reactions that collectively put the aquatic ecosystems to risk [2]. Due to its high toxicity at low concentration, Cd is considered as an important contaminant of natural waters [164]. Research regarding the adverse effects of Cd on microor‐ ganisms demonstrated that Cd2+, via a variation of mechanisms, affected several biochemical algal processes. References include the displacement of Zn2+and Ca2+ co‐factors from unde‐ fined protein targets or directly binding amino acid residues, including cysteine, glutamate, aspartate, and histidine [165]; the inhibition of chlorophyll formation and the reduction of both chlorophyll content and Chl a/b ratio through disturbances in the electron transport chain in both PSI and PSII; and the reduction of Rubisco and enhancement of lipoxygenase activity [2, 145].

Chromium is a transition element that comprises the seventh most abundant metal in the earth's crust, whereas trivalent (Cr3+) and hexavalent (Cr6+) ions are its two most common and stable oxidation states in the environment. Whereas Cr3+ is considered a micronutri‐ ent, essential for the proper function of living organisms, Cr6+ instead can display numerous toxic effects on biological systems. Cr6+ is usually associated with oxygen to form chromate (CrO<sup>4</sup> 2−) or dichromate (Cr<sup>2</sup> O7 2−) oxyanions that can easily go through cell membranes as an alternative substrate for the sulfate transport system and exhibit strong oxidative potential [166]. Therefore, Cr6+ is associated with several intracellular and ultra-structural modifica‐ tions, among which the inhibition of photosynthesis is included. As observed in the cases of the algal species *Chlamydomonas* [134], *C. pyrenoidosa* [167], *Eudorina unicocca*, *C. Kessleri* [168], *E. gracilis* [150], *S. obliquus* [169], and *Monoraphidium convolutum* [170], Cr6+ caused an enhanced destruction of the reaction centers and a reduction in measured Chl-α-fluorescence parameters such as *Φ*PSII, *F*<sup>v</sup> */F*m, *Φ'*PSII, *ETR*, and *qP* [2].

Zinc is an essential element for the activity of several enzymatic systems of organisms. Stimulatory effects on algal photosynthesis at low exposure concentrations of Zn2+ have been observed. For example, *C. vulgaris* after 96 h of exposure at treatment concentration of 5 μmol L−1 showed that the proportion of the maximum quantum yield of PSII promoted by Zn was approximately 10% [154]. However, when the external concentration of Zn2+ is beyond a limited value, it causes harmful effects; hence, its concentration in the cells must be controlled. Zn deficiency in *E. gracilis* has been shown to affect growth, morphology, cell cycle, and mitosis. These observations are best explained by a role for zinc in gene regulation, through zinc-dependent enzymes [149]. Significant effect on *F*<sup>v</sup> */F*m ratio of *P. lanceolatum* (Brébisson) at a concentration level of 0.2 mg L−1 of Zn2+ was observed, while the sensitivity of the same algal species toward all tested heavy met‐ als was diminishing in the order: Cd2+ > Zn<sup>2</sup> > Cu2+ [155].

Copper is unquestionably an essential element in various metabolic processes of algae, such as amine oxidase and cytochrome c oxidase system, prosthetic group of the chloroplastic antioxidant enzyme Cu/Zn superoxide dismutase, and regulator of PSII-mediated electron transport. However, Cu is still considered as one of the most toxic heavy metal ions to algae and is a potent inhibitor of photosynthesis [2]. Many studies have examined ecotoxicological effects of Cu on photosynthetic activity of plants and phytoplankton [145]. From an evaluation of the literature, Cu can affect photosynthetic electron transport on the reducing side of PSI at the level of the ferredoxin [146], alter the PSII on the oxidizing side by inhibiting the electron transport at P680 (the primary donor of PSII) or by inactivating some PSII reaction centers [147]. Cu may also impair the PSII electron transport on its reducing side by affecting the rate of oxidoreduction [148]. The inhibitory effect of copper on the photosynthetic apparatus of several species of algae has been examined, including *E. gracilis* [149, 150], *S. quadricauda* [141], *S. obliquus* [151], *S. incrassatulus* [152], *C. pyrenoidosa* [153], *C. vulgaris* [154], *Planothidium lan‐ ceolatum* and *Isochrysis galbana* [155], *D. tertiolecta, Promocentrum minimum, Synechococcus* sp*.,*

**Table 3.** Examples of metals toxicity on the photosynthetic apparatus reported for various algae. Reports in chronological

of PSII.

**Metallic form Test species Observed stress response References**

Significant effect on *F*<sup>v</sup>

membrane.

Reduction of *F*<sup>v</sup>

activity.

content.

concentrations of Cd2+ ≥0.1, Zn2+ ≥0.2, and Cu2+ ≥0.4 mg L−1.

Modification of mitochondrial

Reduction of photosynthetic

Reduction of the algal *F*<sup>v</sup>

in all tested algae.

both metals. Cu2+ induced the synthesis of chl‐a in *G. edulis* and *G. salicornia* but inhibited chl-α synthesis in *G. manilaensis.* Pb2+ induced the production of Chl-α

content (Chl-α and car) and photosynthetic efficiency (*F*<sup>v</sup>

*/F*m at

*/F*m and Chl-α

*/F*m in

*/F*m)

Sbihi et al. (2012) [155]

Machado et al. (2015) [161]

Bakar et al. (2015) [133]

Bakar et al. (2015) [162]

Deep et al. (2016) [163]

Cd2+, Cu2+, Zn2+ *Planothidium lanceolatum*

Cd2+, Cr6+, Cu2+, Zn2+ *Pseudokirchneriella* 

Hg2+ *Gracilaria salicornia*

Cu2+, Pb2+ *Gracilaria edulis*

order.

58 Chlorophyll

(Brébisson)

*subcapitata*

*Sargassum* sp. *Ulva reticulata*

*Gracilaria manilaensis Gracilaria salicornia*

Pb2+ *Anabaena* sp. Reduction of pigment

Note: *Pseudokirchneriella subcapitata*, known as *Selenastrum capricornutum*.

Cadmium is a heavy metal that occurs naturally in ores along with zinc, lead, and copper. Its compounds are used as stabilizers in PVC products, color pigment, several alloys, and in rechargeable nickel-cadmium batteries. Cd forms complexes with various organic particles and thereby triggers a wide range of reactions that collectively put the aquatic ecosystems

and *Thalassiosira weissflogii* [145].

The toxicity of ionic silver to a variety of aquatic organisms, such as algae, has been studied and shown to be significant, whereas from an evaluation of the literature, Ag+ displayed tox‐ icity to aquatic photosynthetic microorganisms in the nanomolar (nM) concentration range [157, 159]. The toxicity of other forms of silver, such as silver nanoparticles (AgNP) ranged in size from 10 to 200 nm, has been examined as well and according to fluorometry valuesAgNPs were found to influence the photosynthesis of *C. reinhardtii* as well as ionic silver (Ag<sup>+</sup> ) [159].

At this point, it must be mentioned that due to the fact that aquatic ecosystems act as reservoirs of several mixtures of metals, it is essential to evaluate the combined or cumulative effect of metals or metal mixtures on photosynthesis. Therefore, toxicological studies dealing with heavy metal pollution in aquatic organisms must take into account the interactions among metals that may influence uptake, accumulation, and toxicity [2, 128]. For instance, it has been reported that interactions between Cu2+ and Mg2+ may have special significance regarding phytoplankton growth [2]. In another survey assessing the effect of Cu2+, Cr6+, and Ni2+ on growth, photosynthesis, and chlorophyll, a synthesis of *C. pyrenoidosa,* it was demonstrated that various bimetallic combinations of those metals interacted synergistically [171]. Combined effects of Cu2+ and Cd2+ on the growth and photosynthesis-related gene transcription of *C. vulgaris* have been also investigated [154].

In a more realistic approach, metals could also occur along with other contaminants in mixtures. In that respect, synergistic interactions have been predicted between pesticides that act as PSII inhibitors (and are included in the database of Belden et al. [98]) and the metals Cd, Cu, and Zn [101]. A proposed synergistic mechanism between metals and PSII inhibitors in autotrophs could be that metals might prevent the repair of not only damaged PSII com‐ plexes, which are constantly repaired during photosynthesis, but also the damage caused by the reactive oxygen species (ROS) created by the PSII inhibition and the metals themselves, by interacting with enzymes responsible for the repair [101].

Finally, metal bioassays must take into account the synthetic organometallic compounds or the ones formed under environmental conditions. These organometallic substances, especially of Hg, Pb, and Sn, might have completely different toxicological properties and can be more toxic to aquatic organisms because of their high bioaccumulation, as is the cases of methyl mercury compounds (methylation process is thought to be bacterially mediated) [128, 143] and tributyltin chloride [3, 105].

However, it must be underlined that several metal-tolerant algal strains, which have been adapted to environments contaminated with toxic metals (such as Cu and Cd), have been isolated and identified and a variety of tolerance mechanisms have been described [172]. Metallothioneins (MTs) consist one of the most important cellular defense mechanisms against metal stress that regulate the toxicity of various metals and trace elements. MT is a family of cysteine-rich and low-molecular-weight proteins localized to the membrane of the Golgi apparatus, which have the ability to bind several metals through the thiol clusters of their cysteine residues [173]. Some algal MTs are gene products, while others are secondary metabolites [172]. According to rel‐ evant studies, these molecules chelate toxic trace metals, for example, Cd, thereby reducing the concentration of cytotoxic, free-metal ions. Furthermore, some MTs are believed to be involved in zinc and copper homoeostasis [172]. The removal of heavy metals from polluted waters by the use of algae (e.g., *C. pyrenoidosa* and *Scenedesmus* sp.) is called phycoremediation and is an expanding technology with several advantages over physical remediation methods [174].

#### **8. Conclusions and trends**

One of the common and main goals of environmental science and ecotoxicology is the envi‐ ronmental sustainability that concerns the natural aquatic ecosystems and how they endure and remain diverse and productive. Taking into account that photosynthetic microorganisms are the main primary producers and consist of the basis of the food chains, a large number of toxicity tests have been conducted in order to assess the effects of a variety of environmen‐ tal pollutants on algal photosynthetic activity. According to the available vast information, several bioassays have been performed with a great variety of standard test species of both freshwater and saltwater algae, though various "non-standard" algal species have been used on occasion. In our knowledge, in most cases freshwater microalgae were used more fre‐ quently in laboratory toxicity tests than any other types of aquatic plant, except in the case of oil spills where more data for marine algae are available. Moreover, literature data showed that the most commonly used microalgae in marine toxicity tests are green algae and diatoms. The observed differences in response and sensitivity by various microalgal and macroalgal species to the same toxicant can be several orders of magnitude for toxicants such as crude oils, oil products, pesticides, antifouling biocides, and metals. Evidenced heterogeneous sen‐ sitivity of different algal species to the same pollutant is attributed to several characteristics of the exposed alga such as photosynthetic capacity and pigment type, cellular lipid and protein content, and cell size.

of metals or metal mixtures on photosynthesis. Therefore, toxicological studies dealing with heavy metal pollution in aquatic organisms must take into account the interactions among metals that may influence uptake, accumulation, and toxicity [2, 128]. For instance, it has been reported that interactions between Cu2+ and Mg2+ may have special significance regarding phytoplankton growth [2]. In another survey assessing the effect of Cu2+, Cr6+, and Ni2+ on growth, photosynthesis, and chlorophyll, a synthesis of *C. pyrenoidosa,* it was demonstrated that various bimetallic combinations of those metals interacted synergistically [171]. Combined effects of Cu2+ and Cd2+ on the growth and photosynthesis-related gene transcription of

In a more realistic approach, metals could also occur along with other contaminants in mixtures. In that respect, synergistic interactions have been predicted between pesticides that act as PSII inhibitors (and are included in the database of Belden et al. [98]) and the metals Cd, Cu, and Zn [101]. A proposed synergistic mechanism between metals and PSII inhibitors in autotrophs could be that metals might prevent the repair of not only damaged PSII com‐ plexes, which are constantly repaired during photosynthesis, but also the damage caused by the reactive oxygen species (ROS) created by the PSII inhibition and the metals themselves, by

Finally, metal bioassays must take into account the synthetic organometallic compounds or the ones formed under environmental conditions. These organometallic substances, especially of Hg, Pb, and Sn, might have completely different toxicological properties and can be more toxic to aquatic organisms because of their high bioaccumulation, as is the cases of methyl mercury compounds (methylation process is thought to be bacterially mediated) [128, 143]

However, it must be underlined that several metal-tolerant algal strains, which have been adapted to environments contaminated with toxic metals (such as Cu and Cd), have been isolated and identified and a variety of tolerance mechanisms have been described [172]. Metallothioneins (MTs) consist one of the most important cellular defense mechanisms against metal stress that regulate the toxicity of various metals and trace elements. MT is a family of cysteine-rich and low-molecular-weight proteins localized to the membrane of the Golgi apparatus, which have the ability to bind several metals through the thiol clusters of their cysteine residues [173]. Some algal MTs are gene products, while others are secondary metabolites [172]. According to rel‐ evant studies, these molecules chelate toxic trace metals, for example, Cd, thereby reducing the concentration of cytotoxic, free-metal ions. Furthermore, some MTs are believed to be involved in zinc and copper homoeostasis [172]. The removal of heavy metals from polluted waters by the use of algae (e.g., *C. pyrenoidosa* and *Scenedesmus* sp.) is called phycoremediation and is an expanding technology with several advantages over physical remediation methods [174].

One of the common and main goals of environmental science and ecotoxicology is the envi‐ ronmental sustainability that concerns the natural aquatic ecosystems and how they endure

*C. vulgaris* have been also investigated [154].

60 Chlorophyll

and tributyltin chloride [3, 105].

**8. Conclusions and trends**

interacting with enzymes responsible for the repair [101].

Algae have been suggested and used as potential bioindicators of aquatic pollution [1, 175]. Damage of their photosynthetic apparatus is a very sensitive response to xenobiotics that could point to an important biomarker [79]. Carried out studies confirmed that inhibition of photosynthesis is one basic reflex of the toxic effects of several organic and inorganic pollutants on microalgae which in many cases is a more sensitive end point than inhibition of growth [39]. Therefore, we can conclude that measuring the photosynthetic activity is a good screening method for detecting a variety of possible stress situations [132].

Loadings of several anthropogenic pollutants are usually nearly and chronically synchronous with discharges, leading to marked changes in exposure levels of inhabitants of aquatic res‐ ervoirs. Depending on the nature, concentration, frequency, and duration of toxicants expo‐ sure, their impacts on biological communities can prove highly variable [89]. Until nowadays, many experimental studies of aquatic communities of microorganisms have been done using water-column phytoplanktonic species, but only a few have attempted to assess the effect of environmentally realistic pollution exposure scenarios on microbenthic periphyton [89, 122, 123, 176]. The distribution characteristics of chemical toxicants between water phase and sedi‐ ment are of major importance in the evaluation of their fate and ecotoxicological effects into environmental compartments, especially for organic hydrophobic pollutants. Therefore, more vivid studies need to be performed in the future on the bioavailability of organic pollutants and the possible link between pollutant dynamics in the adsorbed phase (bottom sediment periphyton matrices) and their impacts on microbenthic photosynthetic algae.

Last but not least, there is still not much known about the possible toxic effects of transforma‐ tion and degradation products of several synthetic organic compounds on aquatic microal‐ gae. This lack of data makes the toxicity assessment of formed organic molecules metabolites essential, because these molecules may be more toxic than the parent ones; hence, further studies are required to evaluate the adverse effects of these produced chemical species on algal photosynthetic activity.

#### **Author details**

Andreas S. Petsas1,2\* and Maria C. Vagi<sup>2</sup>

\*Address all correspondence to: apetsas@env.aegean.gr

1 Department of Food Science and Nutrition, School of Environment, University of the Aegean, Myrina, Lemnos, Greece

2 Laboratory of Chemical Processes & Aquatic Toxicology, Department of Marine Sciences, School of Environment, University of the Aegean, University Hill, Mytilene, Lesvos, Greece

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