Keywords

1 Introduction

Nanomaterials (NMs), defined as materials with one or more dimensions of the size of 1–100 nm (ASTM/E2456-06 2012; Reiners 2013), have gained increasing attention due to their unique properties, relative to their bulk counterparts, which impart them beneficial characteristics including a high specific surface area and reaction activity (Laurent et al. 2008) and the quantum confinement effects (Amelia et al. 2012). Nowadays with the increasing insertion of nanotechnology in our daily life, nanomaterials of different shapes and diameters have been developed and used in various consumer products, pharmaceuticals, cosmetics (Melo et al. 2015), and other commodities (Bradley et al. 2011; Ghasemzadeh et al. 2014; Singh 2017).

Nanoparticles (NPs), a subgroup of nanomaterials, are classified into various categories depending on their size, morphology, and chemical properties. The primary focus of this chapter is metal, metal oxide, carbon-based NPs, and quantum dots. Metal NPs, i.e., Cu, Ag, and Au, in nanometer range possess unique optical, electrical, and magnetic properties due to their localized surface plasmon resonance (LSPR) characteristics (Dreaden et al. 2012). They are considered as the potential candidates for catalysis due to their large surface area per volume or weight unit, compared to their bulk counterparts, typically functioning on metal surfaces (Roldan Cuenya 2010). Metal oxide nanoparticles represent a class of engineered nanomaterials that can be synthesized via several routes (Lang et al. 2011). The usual practices for manufacturing metal oxide nanoparticles are chemical vapor synthesis (Stankic et al. 2016) and the addition of oxidizing/precipitating agents during their synthesis (Sanchez-Dominguez et al. 2009). Metal oxide nanoparticles include both individual (e.g., TiO2, CeO2, CrO2, ZnO, Bi2O3, and MoO3) and binary oxides (e.g., BaTiO2, InSnO, and LiCoO2). They too have wide applications in industry.

Carbon-based NPs comprise mainly of two major groups: fullerenes and carbon nanotubes (CNTs). Fullerenes are globular cage-like structures with various numbers of carbon atoms (e.g., C60, C70). They have created noteworthy commercial interest due to their remarkable electronic and structural properties, high strength, and versatility (Astefanei et al. 2015). The fullerene-C60 is the most commercially attractive carbon-based NP due to its ability to increase the efficiency of drugs, cosmetics, and electronics (Bianco and Da Ros 2011). Carbon nanotubes (CNTs) have been listed at the third position of the most important engineered nanoparticles (ENPs) found in the consumer product inventories (Vance et al. 2015). They are elongated, tubular structure with a large length/diameter ratio (Ibrahim 2013). They can be metallic or semiconducting reliant on a chiral vector value (the way they are rolled up) (Aqel et al. 2012). Structurally, they resemble graphite sheet rolling upon itself. Depending on the number of carbon sheets, they can either be single-walled (SWNTs), double-walled (DWNTs), or multiwalled carbon nanotubes (MWNTs), respectively (Elliott et al. 2013). CNTs have many potential applications, e.g., in plastics, batteries, paints, composites, touch screens, and drug delivery (De Volder et al. 2013).

Semiconductor materials possess properties between metals and nonmetals and therefore have found various applications (Ali et al. 2017). Semiconductor NPs or quantum dots (QDs) possess wide bandgaps and therefore show significant alteration in their properties with bandgap tuning by exhibiting particle size-dependent tunable photoluminescence (Dybiec et al. 2007). These properties render them very useful materials in photocatalysis, photo optics, electronic devices (Chow and Jahnke 2013), biology, and medicine (Zhou et al. 2015). The most common commercially available QDs are CdSe/ZnS-QDs due to their bright and unique emission with wide excitation spectra and narrow emission bandwidths (Deerinck 2008).

The increased use of nanomaterials in several industrial applications and consumer products ranging from cosmetics to medicine (e.g., odor-resistant textiles, household appliances, including wound dressings) (Rai et al. 2009; Namasivayam et al. 2010; Chaudhari et al. 2012) during the past decade has led to a rise in concerns about the potential toxic effects of accidentally or incidentally released NPs into the environment and their likely access into ambient aquatic systems (Colvin 2003; Service 2008). Scientists have expressed concerns about the potential adverse effects of NPs to beneficial bacteria in the environment, especially in soil and water. Although toxic effects of NPs on bacterial, fungal, and mammalian cells have been well investigated (Shrivastava et al. 2007, 2009; Kim et al. 2009), their impact on the growth and biology of algae and nonvascular lower plants has not been sufficiently documented (Lee et al. 2005).

“Lower plants ” represent a heterogeneous group of plants and plantlike organisms, including algae, bryophytes, and lichens, which are characterized primarily by their lack of vascular tissues (which circulate water and nutrients in higher plants) (Eddy et al. 1992). However, pteridophytes, which are also included in lower plants due to absence of seeds, have the vasculature (phloem and xylem tissues), like the seed-bearing “higher plants.” This chapter is focused on algae and bryophytes, the nonvascular lower plants, because there is hardly any information on responses of sister groups (like lichens, pteridophytes) to NPs.

Algae represent a large and diverse group of photosynthetic eukaryotes. They comprise of many ancient and miscellaneous lineages, including various symbiotic relationships with animals and fungi. Moreover, they display many degrees of organismal complexity: they range from microscopic unicells to macroscopic bodies and also possess multicellular thalli more than a meter in length (De Clerck et al. 2012). Further, algal cell wall surface has an additional layer of rigid, porous cell wall for modulating the entry of foreign materials, ions, and particles (Chen et al. 2012). Algae are commercially important since they can be used as biofertilizers, pollution control agents (algae bioreactors) (Pimratcha et al. 2015), biofuels (Hannon et al. 2010), stabilizers of casein, and source of nutrition (B complex vitamins and minerals) and can be incorporated to cosmetics (Spolaore et al. 2006). They form a critical component of almost all aquatic and many terrestrial ecosystems. As primary producers in the aquatic ecosystem, they are important indicators for environmental pollution monitoring and therefore constitute widely used model organisms in ecotoxicity studies of nanomaterials (Ma and Lin 2013; Quigg et al. 2013). The observed toxicities of NPs to algae have been attributed mainly to three mechanisms (Schwab et al. 2011; Long et al. 2012): (1) reduction of photosynthetic rate due to inhibition of light transmittance (shading effect) (Miazek et al. 2015); (2) NP agglomeration and physical interaction with algal cells, leading to the internalization of NPs and the disruption of the cell membrane (García-Cambero et al. 2013); and (3) induction of intracellular reactive oxygen species (ROS) formation leading to membrane lipid peroxidation and changes in the concentration of nonenzymatic antioxidants and in the activity of antioxidant enzymes (Chen et al. 2012).

Like algae, bryophytes are also nonvascular plants that can grow on the surface of tree trunks or rocks and generally absorb water and nutrients direct through leaf surfaces from their immediate environment (Dymytrova 2009; Schröder et al. 2010; Harmens et al. 2011). Some species play an important role in the colonization of bare or degraded soil and facilitate the installation and maintenance of higher plants, a significant fact for a healthy plant-soil dynamics and sustainable ecosystem development. Moreover, these lower plants are recognized as good accumulators of pollutants, especially for metal trace elements (Garrec and Van Haluwyn 2002; Faburé et al. 2010). Furthermore, these organisms lack a root system ensuring their exposure to atmospheric pollutants and are excellent sensors of air quality for different contexts and pollutants (Faburé et al. 2010; Meyer et al. 2010; Lodenius 2013). Owing to these properties, bryophytes are considered excellent models for evaluating atmospheric pollutant impact on the environment in many parts of the world (Oishi 2013; Agnan et al. 2015; Vuković et al. 2015). The observed toxicities of NPs to bryophytes have been mainly attributed to (a) NP agglomeration that leads to the NP uptake by leaves (Canivet et al. 2014) and (b) overproduction of ROS/RNS and induction of glutathione status modulation (Canivet et al. 2015).

Since the last decade, the use of NPs in our daily life has considerably increased, and the ecotoxicity studies about NP effects on plants and animals come up rapidly. However, the published data on NP toxicity on algae and plants of sister groups are insufficient, and the experimental designs and testing conditions are inconsistent across the studies. This chapter provides an outline of the NP ecotoxicological effects on these plants based on the data available in the published literature and focuses on the underlying mechanisms of NP toxicity.

2 Effects of Metal Nanoparticles

2.1 Gold Nanoparticles (Au-NPs)

A recent study explored the impacts of amine-coated 10 nm gold NP contaminations on the green alga Scenedesmus subspicatus (Renault et al. 2008). The growth/mortality effects were determined by the algal cell numerations. The lethal dose for 50% of the population was reached within a 24 h exposure at 1.6 × 105 Au-NPs/cell, while mortality for the lowest contamination condition was 20%. TEM examination revealed that Au-NPs were strongly adsorbed by the cell wall of algae, leading to progressive intracellular and wall disturbances. However, bare and hyaluronic acid-capped Au-NPs (particle size 12.5 nm) were found to be harmless to S. subspicatus in contrast to bulk soluble gold that gave an EC50 value 1.91 mg L−1 (García-Cambero et al. 2013). The carbonate- and citrate-coated Au-NPs did not cause any significant toxicity to the green algae Chlamydomonas reinhardtii (Behra et al. 2015).

The effects of Au-NPs with amphiphilic polymer coating (AP) or amphiphilic coating to which 10 kDa polyethylene glycol chains were attached (AP-PEG) were also assessed on the green algae Pseudokirchneriella subcapitata (Van Hoecke et al. 2013). The AP Au-NPs were more toxic than the AP-PEG after a 72 h exposure with EC50 values 7.5 and 39 mg Au-NPs L−1, respectively. In another study Au-NP toxicity effects on P. subcapitata (formerly known as Selenastrum capricornutum) were evaluated by three biomass measuring techniques (coulter counting, cell counting in hemocytometer, and fluorescence of pigment extracts) (Hartmann et al. 2013). The coulter counting method gave unreliable results. Therefore, it was not considered suitable for biomass quantification. At 48 h, algal cultures showed follow the growth of the control sample – both with respect to cell number and pigment content – whereas at 72 h a leveling off was noticed in pigment content of all algal cultures exposed to Au-NP dispersions accompanied by a decrease in EC10. On the contrary, cell number growth rates followed continued exponential trends. The results also indicated that pigment (such as chlorophyll and carotenoid) synthesis was affected by the exposure to Au-NP dispersions despite the continued exponential cell growth. Although the EC50 values were higher than the highest tested concentration, subsequent tests of the effects of the “Starch Control” (starch/glucose/MES solution in concentrations identical to that of the Au nanoparticle dispersions) revealed that the dispersion constituents, and not Au-NPs themselves, were largely responsible for inhibitory effects as well as the characteristic leveling off in pigment content after 48 h.

A recent study suggested that the naturally existing ions like zinc ions could modulate the toxicity of Au-NPs (Iswarya et al. 2017). The effects of Au-NPs with two surface cappings (citrate and PVP) and three different sizes (16, 27, and 37 nm) were explored on a predominant freshwater alga Scenedesmus obliquus in the sterile freshwater matrix. Among the different-sized Au-NPs, the highest toxicity (54%) was observed at 1 mg L−1 of citrate-capped Au-NPs with particle size 37 nm, whereas PVP-capped Au-NPs showed 42% toxicity. A statistically significant reduction in the Au-NP (both citrate-capped/37 nm and PVP-capped/37 nm) toxicity was observed when Zn2+ (5 mg L−1) was added to the growth medium (12% for citrate-capped −37 and 11% for of PVP-capped −37). All the above information on the Au-NP toxicity on lower plants is summarized in Table 16.1, specifying the NP size, the algal species used, the half maximal effective concentration, the exposure time, and the effects observed.

Table 16.1 Summarized results from Au-NP toxicity studies performed on the lower plants
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2.2 Silver Nanoparticles (Ag-NPs)

Ag-NPs have shown extensive adverse effects on growth and morphology of the green algae Pithophora oedogonia and Chara vulgaris in a dose-dependent manner (Dash et al. 2012). Exposure of algal thalli to increasing concentrations of Ag-NPs resulted in progressive chromosome instability, mitotic disturbance, depletion of chlorophyll content, and the associated morphological malformations in algal filaments. SEM micrographs revealed dramatic alterations in cell wall, characterized with cell wall rupture and degradation in the NP-treated Pithophora. Discoloration of filaments due to chloroplast contraction followed by disintegration, regional bulging of filaments, thinning and disruption of cell wall permitting exclusion of the chlorophyll pigments, adsorption of Ag-NPs on cell surface and organellar membranes, mitostatic effect, induction of chromosomal anomalies, and irreversible genetic damage were the significant detrimental effects of nanosilver recorded in the tested algae. Ag-NPs also caused growth inhibition on the green algaPseudokirchneriella subcapitata with an EC50 value 0.19 mg Ag-NPs L−1 after 96 h incubation (Griffitt et al. 2008).

The short-term toxicity of citrate-stabilized Ag-NPs and ionic silver Ag(I) to the ichthyotoxic marine raphidophyte Chattonella marina has also been investigated (He et al. 2012). The addition of Ag-NPs to GSe medium caused aggregation and dissolution of Ag-NPs. Cellular uptake of dissolved Ag(I) was observed, and toxicity effects were much higher for Ag(I) than for Ag-NPs. However, these inhibitory effects of Ag(I) and Ag-NPs were completely removed by the addition of cysteine, a strong Ag(I) ligand, suggesting that the toxicity of Ag-NPs was due to the release of Ag(I). The growth inhibition effects of Ag-NPs have been studied also on the bloom-forming cyanobacterial Microcystis aeruginosa strain after a 10-day exposure (Duong et al. 2016). A dose-dependent reduction of the cell growth was observed by increasing Ag-NP concentrations. The EC50 value based on the cell growth was 0.0075 mg L−1, and the inhibition efficiency at the highest concentration of Ag-NPs (1 mg L−1) was 98.7%. SEM and TEM images indicated shrunken and damaged cell wall attributed to toxicity of Ag-NPs.

The toxic effects of large-sized Ag-NPs (50 nm) were investigated on the freshwater microalga Chlorella vulgaris and the marine microalga Dunaliella tertiolecta after 24 h exposure (Oukarroum et al. 2012). Ag-NPs interacted directly with the Chlorella vulgaris cell surface forming large aggregates and caused significant decrease in chlorophyll contents and algal viability, while they induced ROS formation and lipid peroxidation in both algae, showing a variability in sensitivity (1 mg L−1 Ag-NPs induced a 44% decrease of viable cells for D. tertiolecta and 33% for C. vulgaris). In another study, aggregation and dissolution behavior of gum arabic (GA)- and polyvinylpyrrolidone (PVP)-coated Ag-NPs were compared in a mixture of aquatic plants Potamogeton diversifolius and Egeria densa (Unrine et al. 2012). Plants released dissolved organic matter (DOM) into the water column either through active or passive processes in response to Ag exposure that bound Ag ions. As a result, the plant-derived DOM stabilized PVP-Ag-NPs as the primary particles but removed GA-Ag-NPs from the water column, possibly by dissolution and binding of the released Ag ions on sediment and plant surfaces.

The extent and mechanisms of toxicity of two Ag-NPs with differing size distributions (AG1 and AG2) and capping agents were investigated on two model organisms, a green alga (Chlamydomonas reinhardtii) and a cyanobacterium (Synechococcus leopoliensis) (Taylor et al. 2016a). Their effects on the production of extracellular polymeric substances (EPS) were also assessed. Both silver forms had a significant effect on viability and membrane integrity in C. reinhardtii but hardly affected ROS production, whereas no toxicity effects were observed in S. leopoliensis. The levels of EPS produced by both the species were similar for all the treatments. The EPS composition was affected from AG1 in a concentration-dependent manner and conversely from AG2. Higher levels of lower molecular weight material were produced by C. reinhardtii in the presence of all silver forms. Reduction in growth rate was observed for S. leopoliensis, but the impact on viability and ROS was lower than for C. reinhardtii probably due to differences in relevant biological properties (e.g., algal cell size and cell wall composition).

The abovementioned toxicity data of Ag-NPs on lower plants are presented in Table 16.2, specifying the particle size of Ag-NPs, the algal species used, the half maximal effective concentration and exposure time, and the effects observed.

Table 16.2 Summarized results from Ag-NP toxicity studies performed on lower plants
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2.3 Platinum Nanoparticles (Pt-NPs)

Pt-NP toxicity toward green microalgae Pseudokirchneriella subcapitata and Chlamydomonas reinhardtii was assessed (Sørensen et al. 2016), using the standard ISO tests for estimation of growth rate inhibition (EC50 values of 15−200 mg Pt-NPs L−1). By using a double-vial setup, cells were separated from Pt-NPs, which indicated that shading is an important artifact for Pt-NP toxicity. Membrane damage was not severe, but substantial oxidative stress was detected at 0.1–80 mg Pt-NPs L−1 in both the algal species. Pt-NPs caused a growth rate inhibition and oxidative stress in P. subcapitata, in a low concentration of dissolved Pt, indicating the NP-specific toxicity of Pt. In addition, higher body burdens were measured in this species, possibly due to a favored binding of Pt to the polysaccharide-rich cell wall. However, in a previous study, the Pt-NP concentration causing total inhibition of algal growth was 22.2 mg L−1 (Książyk et al. 2015). Similar results were obtained by analyzing the levels of photosynthetic pigments in P. subcapitata exposed to nanoparticles. In another study, where the acute toxicity of PtCl4 and Pt-NPs was investigated, EC50 values were 14 mg L−1 after 48 h exposure and 28 mg L−1 after 2 h, indicating that the toxicity was dependent on the exposure duration (Delgado et al. 2013).

3 Effects of Metal Oxide Nanoparticles

3.1 Alumina Nanoparticles

The toxicological impact of Al2O3-NPs to lower plants was demonstrated on algal species Scenedesmus sp. and Chlorella sp. (Aruoja et al. 2015). The observed EC50 value of Al2O3-NPs with particle size <50 nm after 72 h was 45.4 mg L−1 for Chlorella sp. and 39.35 mg L−1 for Scenedesmus sp. Bulk alumina (particle size <5 um) also showed toxicity in a lower range (EC50 = 110.2 mg L−1 for Chlorella sp.; 100.4 mg L−1 for Scenedesmus sp.). Additionally, chlorophyll content declined, and the cell surface was also affected (Sadiq et al. 2011b). Earlier, Al2O3-NPs with a particle size of 51 nm showed an EC50 value 8.30 mg L−1 at 96 h for the algal species Pseudokirchneriella subcapitata (Griffitt et al. 2008), whereas smaller Al2O3-NPs (particle size 8–21 nm) showed an EC50 value >10–100 mg Al2O3 L−1 at 72 h.

In addition, toxic effects of binary compounds of aluminum oxide alpha-forms (7 and 70 nm) and macro form (4um) on the growth of unicellular algae Chlorella vulgaris (Gosteva et al. 2015) were found to be concentration-dependent. A selective dependence of Al2O3-NPs toxicity on the size, concentration, and chemical nature of NPs was revealed. The EС50 values for the small-sized Al2O3-NPs (7 and 70 nm) were around 1 mg L−1, whereas no toxicity was observed for the macro form.

3.2 Cerium Oxide Nanoparticles (CeO2-NPs)

An assessment of the molecular and phenotypic effects of CeO2-NPs was conducted with the unicellular green alga, Chlamydomonas reinhardtii, by using well-characterized monodispersed NPs (particle size 4– nm) based on the hypothesis that their toxicity is likely to be higher than the macro form (Taylor et al. 2016a, b). The potential toxicity of NPs was investigated by transcriptomics and metabolomics approaches in a wide range of exposure concentrations in order to provide insight into molecular toxicity pathways. Even though CeO2-NPs inserted the intracellular vesicles within C. reinhardtii, they did not cause significant changes in the algal growth at any exposure concentration. At supra-environmental CeO2-NPs concentrations, downregulation of photosynthesis and the carbon fixation perturbations were detected with further effects on energy metabolism.

In a previous study, on a short-term exposure, dissolved Ce3+ decreased the photosynthetic yield in a concentration-dependent manner with EC50 values of 7.5 uM for the wild type and 6.3 uM for a cell wall-free mutant strain of C. reinhardtii, whereas precipitated CePO4(s) was not bioavailable and, hence, not toxic (Röhder et al. 2014). The intracellular ROS levels increased upon exposure to Ce3+ with the effective concentrations being similar to those inhibiting photosynthesis. Moreover, CeO2-NPs agglomerated in exposure media and caused a slight inhibition of photosynthesis and reduction of intracellular ATP content upon a short-term exposure at the highest (100 uM) concentrations possibly due to Ce3+ ions co-occurring in the nanoparticle suspension, whereas no effect was observed for dispersed CeO2-NPs as the dissolved Ce3+ got precipitated with phosphate and, hence, was not bioavailable. Moreover, flocculation of algal cells upon exposure to agglomerated CeO2-NPs and Ce3+ was observed. The cell wall-free mutant and wild type of C. reinhardtii showed the same sensitivity to either Ce3+ or CeO2-NPs toxicity indicating that the cell wall does not have a protective effect against CeO2-NPs or Ce3+.

3.3 Titanium Dioxide Nanoparticles

The studies on a potential impact of TiO2-NPs on the environment have been conducted on green alga Desmodesmus subspicatus by determining its growth during a 72 h incubation period (Hund-Rinke and Simon 2006). Twenty-five nm-sized TiO2-NPs (EC50 = 44 mg TiO2-NPs L−1) were more toxic to D. subspicatus than 100 nm TiO2-NPs (no toxicity observed). It was demonstrated that the smaller particles caused a clear dose-dependent reduction in the algal growth, whereas the larger ones showed less toxicity (C < 50 mg TiO2-NPs L−1). On similar lines, TiO2-NPs (10, 30, 300 nm) were applied to the algal species Pseudokirchneriella subcapitata, and their effects assessed after 72 h incubation (Hartmann et al. 2010). Smaller TiO2-NPs (<10 nm) showed inhibition (21% reduction) in growth rate at the concentration of 2 mg L−1, whereas, 30 and 300 nm TiO2-NPs showed a slight stimulation of algal growth. In earlier studies, TiO2-NPs with particle size of 30 nm (Griffitt et al. 2008) and ~ 100 nm (Blaise et al. 2008) were shown not to be toxic to this alga.

In another study, P. subcapitata was used for toxicity assessment of not readily soluble NPs with standardized algal growth inhibition tests (Hartmann et al. 2013). TiO2-NPs formed large (micron-sized) agglomerates/aggregates in a dose-dependent manner. Three biomass surrogate measuring techniques (coulter counting, cell counting in hemocytometer, and fluorescence of pigment extracts) were evaluated. The results showed a concentration-dependent reduction in algal growth by both the biomass quantification techniques, yielding an EC50 value of 160 mg TiO2-NPs L−1 (by hemocytometer). The EC50 value based on measurements of pigment fluorescence was found to be 200 mg L−1 (the highest tested concentration was 560 TiO2-NPs L−1). P. subcapitata was also used for toxicity assessment of fine (140 nm) and ultrafine (~140 nm) TiO2 (uf-TiO2) particles after incubation for 72 h (Warheit et al. 2007). EC50 values (95% fiducial limits) based on inhibition of growth and healthy average cell counts were 16 mg L−1 for fine TiO2-NPs and 21 mg L−1 for uf-TiO2-NPs. Aruoja et al. (2009) found that bulk TiO2 (EC50 = 35.9 mg Ti L−1) were less toxic to this algal species than their nano formulations (EC50 = 5.83 mg TiO2-NPs L−1). TiO2-NPs formed characteristic aggregates entrapping the algal cells, thus contributing therefore to the toxic effect of TiO2-NPs to algae. In a later study, it was indicated that the agglomerates entrapped nearly all algal cells so that the cells could mostly be seen inside the agglomerates and rarely in the surrounding medium (Aruoja et al. 2015). The high variability in the observed toxicity of TiO2-NPs has been discussed by Menard et al. (2011), but no discernable correlation between primary particle size and toxic effect could be proved because the existing data were insufficient for confirmation (Menard et al. 2011).

Manier et al. (2016) indicated that the type of exposure system also affects the toxicity of TiO2-NPs. Different exposure systems including the Erlenmeyer flasks and 24-well microplates (both using an orbital shake system) and an alternative system using cylindrical vials and magnetic stirring were used. After a 72 h exposure of P. subcapitata to two different types of TiO2-NPs (particle size <10 and 20 nm), the authors found that the exposure systems applied to achieve the test can substantially affect the ecotoxicological results and the subsequent calculated EC50 values. The selected systems influenced both the interaction between algal cells and TiO2-NPs as well as the growth inhibition level (Manier et al. 2016). The cytotoxicity potential of TiO2-NPs was also assessed toward freshwater algal isolate Scenedesmus obliquus under dark and UV conditions at low exposure levels (≤1 μg mL−1) (Dalai et al. 2013). Statistically significant reduction in cell viability and photosynthetic pigment content and increase in ROS production and membrane permeability (light vs. dark) were observed. Cell viability at 1 μg mL−1 concentration under UV illumination and dark conditions was 59.1% and 69.46%, respectively, for 72 h exposure period. Cellular uptake of NPs was indicated in electron micrographs, whereas fluorescence micrographs and images from confocal laser scanning microscopy (CLSM) brought out their probable genotoxic effects (Dalai et al. 2013).

In addition, a comparative study was conducted by Sadiq et al. (2011a, b) to demonstrate the toxic effects caused by TiO2-NPs toward the freshwater algae (Scenedesmus sp. and Chlorella sp.) isolated from freshwater environment after 72 h incubation. The particles had a growth-inhibiting effect for both species (EC50 = 16.12 mg TiO2-NPs L−1 for Chlorella sp. ; EC50 = 21.2 mg TiO2-NPs L−1 for Scenedesmus sp.). Bulk (micron-sized) TiO2 also showed toxicity though to a lesser extent (EC50 = 35.50 mg TiO2-NPs L−1 for Chlorella sp.; EC50 = 44.40 mg TiO2-NPs L−1 for Scenedesmus sp.). A concentration-dependent reduction in the fluorescence of chlorophyll content was also observed (Sadiq et al. 2011a, b). These species were also used in a comparative study of the photocatalytic activity of P25 TiO2-NPs under dark, visible light and UVA conditions (Roy et al. 2016). Chlorella was more sensitive toward the toxicity effects than Scenedesmus. Furthermore, at the highest exposure concentration, ROS generation was found correlated with inactivation of the antioxidant enzymes (SOD and GSH) for both the algae under visible light and UVA conditions. Additionally, TiO2-NPs increased catalase activity and LPO release, indicating the membrane damage, particularly high in Chlorella, which is a single-cell algae and therefore is more susceptible to TiO2-NP uptake in comparison to Scenedesmus, which shows a high colonization tendency.

Moreover, toxicity of NPs of binary compounds of titanium dioxides (with particle size 5, 50, 90, and 350 nm) was studied again on the unicellular alga Chlorella vulgaris (Gosteva et al. 2015). Substantiating the findings of Hartmann et al. (2010), this study revealed a selective dependence of TiO2-NP toxicity on size and concentration of NPs. TiO2-NPs with particle size 5 and 90 nm were classified to the category “acute toxicity 1,” whereas no acute toxicity was registered for particle size 50 nm. Physiological, biochemical, and molecular genetic levels were assessed on the unicellular green algaChlamydomonas reinhardtii after application of TiO2-NPs (Wang et al. 2008). Growth inhibition was observed during the first 2–3 days of incubation with TiO2-NPs, but later a dose-dependent recovery was observed. Oxidative stress occurred within the cell after exposure to TiO2-NPs, which caused an increase in malondialdehyde levels, while four stress response genes (sod1, gpx, cat, and ptox2) were upregulated in cultures containing even 1 mg L−1 of TiO2-NPs. The maximum transcripts of cat, sod1, gpx, and ptox2 occurred at 1.5, 3, 3, and 6 h, respectively, proportional to the initial concentration of the NPs.

Kulacki and Cardinale (2012) examined how TiO2-NPs (ranging from 0 to 300 mg TiO2-NPs L−1) affect the population dynamics and production of biomass across a range of the North American freshwater algae (Anabaena spp., Navicula subminuscula, Nitzschia pusilla, Oscillatoria spp., Planothidium lanceolatum, Scenedesmus quadricauda, Selenastrum minutum, Spirogyra communis, Stigeoclonium tenue, Tabularia fasciculata). The effects of TiO2-NPs on the population growth rate of each algal species over a period of 25 days were not significant (p = 0.376), even though there was a considerable species-specific differentiation in responses (strong inhibition of maximum growth rate for Spirogyra communis, whereas strong stimulation of maximum growth rate for Stigeoclonium tenue). On the contrary, exposure to TiO2-NPs tended to increase the maximum biomass achieved by species in culture (p = 0.06).

Finally, the effects of TiO2-NPs and bulk particles on the marine microalga Nitzschia closterium were evaluated with reference to growth, oxidative stress, and cellular uptake after 96 h incubation (Xia et al. 2015). Toxicity of TiO2-NPs to algal cells significantly increased with decreasing nominal particle size, and the EC50 values were 88.78, 118.80, and 179.05 mg L−1 for 21, 60, and 400 nm NPs, respectively. The growth was significantly inhibited on exposure to 5 mg L−1 of 21 nm TiO2 NPs. Activities of antioxidant enzymes, viz., peroxidase (POD), superoxide dismutase (SOD), and catalase (CAT), were induced at the beginning and thereupon were inhibited, whereas malondialdehyde (MDA) levels and reactive oxygen species (ROS) increased following the exposure to 5 mg L−1 TiO2 NPs, indicating damages on the cell membrane. Flow cytometry and TEM studies and Ti content measurements indicated that TiO2-NPs were internalized in N. closterium cells. The level of extracellular ROS was negligible, as compared to the intracellular ROS level, suggesting that the elevated TiO2 toxicity in marine environments is related to increased ROS levels caused by internalization of TiO2-NPs.

The abovementioned toxicity data of TiO2-NPs are presented in Table 16.3, specifying the particle size, the algal species used, the half maximal effective concentration, and the exposure time used.

Table 16.3 Summarized results from TiO2-NP toxicity studies performed on algae
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3.4 Zinc Oxide Nanoparticles

Toxicity of ZnO-NPs to the algae Pseudokirchneriella subcapitata was determined by the OECD 201 algal growth inhibition test (Aruoja et al. 2009). EC50 values of bulk and nano ZnO particles were both similar to that of ZnSO4 (at 72 h EC50 ~ 0.04 mg ZnO-NPs L−1), and no aggregation formation was observed. This was attributed to the dissolved Zn possibly because most of the ZnO is dissolved at these low concentrations (Franklin et al. 2007). The toxicity data were close to those obtained by Franklin et al. (2007) who showed EC50 values for the same algal species to be 0.063 mg Zn L−1 for bulk ZnO and 0.068 mg Zn L−1 for nano ZnO after 72 h exposure. In another study performed on P. subcapitata, EC50 value was between 0.1 and 1 mg ZnO-NPs L−1 with particle size 8–21 nm (Aruoja et al. 2015).

Toxicity of ZnO-NPs with particle size 20 nm was also evaluated on the unicellular alga Chlorella vulgaris (Morgalev et al. 2015). A concentration-dependent reduction in the fluorescence of growth rate was observed after application of ZnO-NPs. The detected value of EC50 was 0.17 mg ZnO-NPs L−1. In assessing the maximum effect of ZnO-NPs on Chlorella and other organisms according to GHS and EU Directive 93/67/ EEC, they were assigned to dangerous substances with a high-degree toxicity “acute toxicity 1.”

3.5 Iron Oxide and Zerovalent Iron Nanoparticles

Engineered zerovalent nano-iron particles (Fe-NPs) hold promise for remediation of several pollutants, but their impact on the environment is not completely clear. The effects of three types of nZVI, (a) Nanofer 25 (uncoated), (b) Nanofer 25S (surface coated with a Na-acrylic copolymer), and (c) Nanofer STAR (Surface stabilized Transportable Air-stable Reactive) powder with an inorganic coating, were assessed on the growth, cell morphology, and metabolic status of marine microalgae Pavlova lutheri, Isochrysis galbana, and Tetraselmis suecica after 23-day exposure (Kadar et al. 2012). The algal growth rate , size distribution, and cellular structure were not altered significantly in any of the three species. The total cellular lipid content increased in T. suecica grown on media enriched with uncoated Nanofer 25 and in P. lutheri with Nanofer STAR, when compared at equimolar exposures. Furthermore, there occurred a significant change in fatty acid composition complementing the Nanofer STAR-mediated increase in lipid content of P. lutheri. Likewise, Zehnder medium fortified with zerovalent Fe-NPs (Nanofer 25 and Nanofer 25S) boosted the growth of four green algae (Desmodesmus subspicatus, Dunaliella salina, Parachlorella kessleri, and Raphidocelis subcapitata), two eustigmatophycean algae (Nannochloropsis limnetica and Trachydiscus minutus), and the cyanobacterium Arthrospira maxima (Pádrová et al. 2015). In all the species studied, zerovalent Fe-NPs induced lipid accumulation, the saturated and monounsaturated fatty acid (except palmitoleic acid) and polyunsaturated fatty acid contents in cells. The authors suggested that these particles may provide a source of iron that increases cell growth and enhances metabolic changes leading to higher lipid production and changes in the composition of fatty acids.

In another study, toxicities of four zerovalent Fe-NPs of different sizes (20, 50, 80, and 100 nm), Fe2O3-NPs of two sizes (30 and 20 nm) of different crystal phases (α, γ), and Fe3O4-NPs of one size (20 nm) were assessed with green alga Chlorella pyrenoidosa, focusing on the effects of particle size, crystal phase, oxidation state, and environmental aging (Lei et al. 2016). The results indicated a significant increase in toxicity as particle size decreased. The algal growth inhibition decreased with oxidation of the NPs with an order of zerovalent Fe-NPs > Fe3O4 NPs > Fe2O3 NPs, while α-Fe2O3 NPs (EC50 = 71 mg L−1) presented significantly higher toxicity than γ-Fe2O3 NPs (EC50 = 132 mg L−1). The EC50 values after a 96 h exposure to zerovalent Fe-NPs were for 100 nm (91.3 mg L−1) > 80 nm (81.2 mg L−1) > 50 nm (74.1 mg L−1) > 20 nm (19.8 mg L−1). The NP-induced oxidative stress was the main toxic mechanism, which could give a possible explanation in the difference in algal toxicity caused by NPs with the contribution of agglomeration and physical interactions.

The effects of Fe-NPs have also been assessed on the bryophyte Physcomitrella patens subsp. patens after foliar exposure (Canivet et al. 2015). The effects (cytotoxicity, oxidative stress, lipid peroxidation of membrane) of Fe-NPs from industrial emissions of metallurgical industries were determined through the axenic culture of P. patens exposed at five different concentrations (5 ng, 50 ng, 500 ng, 5 mg, and 50 mg per plant). At concentrations tested over a short period (24 h, 72 h), the levels of ROS, MDA, and glutathione were not significantly disturbed, but after internalization (168 h) the Fe-NPs could interact with the intracellular medium and cause cytotoxic effects and/or oxidative stress. Additionally, confocal microscopy experiments revealed that Fe-NPs (particle size 20–80 nm) penetrated the leaves of Aphanorrhegma patens when applied as mineral water suspensions (Canivet et al. 2014). By the way, this was the first demonstration of NP uptake by a bryophyte, and the actual penetration mechanism remains mysterious as also in higher plants.

3.6 Copper Oxide Nanoparticles

In a recent study, toxicity of CuO-NPs toward the algal species Pseudokirchneriella subcapitata was estimated by the OECD 201 algal growth inhibition test (Aruoja et al. 2009). CuO-NPs with mean particle size 30 nm were found to cause higher toxicity (72 h EC50 = 0.71 mg CuO-NPs L−1) than one caused by bulk CuO (72 h EC50 = 11.55 mg CuO-NPs L−1). No aggregates were observed in the growth medium. These findings are in agreement with those of a previous study where 15–45 nm CuO-NPs gave 0.54 mg CuO-NPs L−1 EC50 value at 96 h (Griffitt et al. 2008). The toxicity of CuO-NPs to algae has also been assessed in the presence of dissolved organic matter (DOM). One of the main fractions of DOM (Suwannee river fulvic acid) (20 mg L−1) was added alone or in the presence of CuO-NPs in the culture medium of the prokaryotic alga Microcystis aeruginosa (Wang et al. 2011). Internalization of CuO-NPs was observed in the intact algal cells at certain locations (e.g., thylakoids and granules), and the cell uptake was enhanced by Suwannee river fulvic acid (SRFA). The main form of intracellular NPs observed was Cu2O, indicating that intracellular environment may reduce CuO into Cu2O. The increased CuO nanotoxicity observed in the presence of SRFA was related to the decreased rate of aggregation formation, the higher Cu2+ release, and the induction in the internalization of CuO-NPs.

In another study, the short-term effects of core-shell copper oxide NPs (CS-CuO-NPs) in two different agglomeration states on the green alga Chlamydomonas reinhardtii were examined, and toxicity was investigated with regard to change in cellular population structure, primary photochemistry of photosystem II, and the ROS formation (Saison et al. 2010). CS-CuO-NPs induced cellular aggregation processes and reduced chlorophyll levels by inhibiting photosystem II. This process (inhibition of photosynthetic electron transport) induced a strong energy dissipation process via non-photochemical pathways indicating the formation of reactive oxygen ROS. However, no ROS formation was observed when C. reinhardtii was exposed to the core without the shell or to the shell only. The toxicity of carbon-coated copper nanoparticles (Cu-NPs) was also investigated using the alga C. reinhardtii and compared with effects of dissolved Cu2+ (provided as CuSO4) (Müller et al. 2015). The Cu-NPs agglomerated in the medium from original size of 6–7 nm to average particle sizes of 140–200 nm possibly due to the hydrophobic properties of the carbon coating. Cu-NPs strongly decreased the photosynthetic yield of C. reinhardtii after 1–2 h exposure to dissolved CuII in a concentration range 1–100uM, whereas this decrease occurred in a concentration range of 0.1–10 uM for CuSO4. Cu-NP effects on photosynthetic yield were similar for the same concentration of dissolved Cu2+ for 1 h exposure and slightly stronger after longer exposure times. After the addition of EDTA as a strong ligand for CuII, toxicity of both dissolved CuII and of Cu-NPs was completely suppressed.

The toxicity effects of sonicated and non-sonicated CuO suspensions (<50 nm) were elucidated on macrophytic (Nitellopsis obtusa) and microphytic (Chlorella spp.) algae cells (Manusadžianas et al. 2012). Cell lethality and resting potential depolarization were used to measure the NP effects on N. obtusa, whereas photosynthetic efficiency was assessed on Chlorella spp. There were no substantial differences between the effects of non-sonicated and sonicated CuO-NP suspensions. The particles rapidly reagglomerated within 5 min after sonication. The lethal concentrations of CuO-NPs did not evoke a rapid cell membrane depolarization in N. obtusa within the initial 90 min period, indicating that charophyte cell wall might have delayed the NP toxic effects. Significant cell membrane depolarization could be observed only after a 6 h exposure. In addition, rewash lethality tests revealed that 5 min exposure in 100 mg CuO-NPs L−1 concentration induced algal cells mortality by 70% after 8 days, whereas 6 h exposure at 0.64 mg L−1 of Cu2+ evoked less than 40% cell mortality. The observed lethal effects of algae cells as well as delayed cell membrane depolarization were evoked by nanoparticles or their agglomerates per se, but not by dissolved Cu, as neither chemical analysis nor biological testing confirmed the presence of Cu2+ in toxic amounts.

The macrophyte Lemna gibba was also used to evaluate CuO-NP toxicity. On exposure to CuO-NPs or soluble copper for 48 h, photosynthetic activity was inhibited due to inactivation of photosystem II reaction centers, causing a decrease in electron transport rate and an increase of thermal energy dissipation (Perreault et al. 2014). Toxicity of CuO-NPs was mainly driven by copper ions released from particles due to the NPs’ tendency to agglomerate in the culture medium.

The data on CuO-NPs versus algae are presented in Table 16.4.

Table 16.4 Summarized results from CuO-NP toxicity studies performed on lower plants
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3.7 Nickel Oxide Nanoparticles

NiO-NPs (20 nm average size) were found to provoke a severe growth inhibition on a marine microalga strain of Chlorella vulgaris in a sterilized enriched seawater medium (f/2 medium) when treated with 40–50 mg L−1 during 72–120 h of exposure, with EC50 being 32.28 mg NiO L−1 at 72 h and 44.33 mg NiO L−1 at 120 h (Gong et al. 2011). The observed inhibitory effect was accompanied by cellular structural alterations such as cytomembrane breakage (detached or degraded plasma membrane), plasmolysis (leak of cytosol), and disorder of thylakoids. At the same time, living algae showed a tendency to increase the agglomeration-deposition capacity of NiO-NPs as well as to reduce them for zero valence nickel.

In another study, the aquatic plant Lemna gibba was used to investigate and compare the toxicity induced by 30 nm NiO-NPs and nickel(II) oxide as bulk (NiO Bulk) (Oukarroum et al. 2015). Plants were exposed for 24 h to NiO-NPs, or NiO Bulk caused agglomerations of NiO-NPs in culture medium, due to ionic strength. Both NPs and bulk enhanced ROS formation, especially at 1000 mg L−1 (five times compared to control), indicating the cellular oxidative stress. Both types of NiO induced a strong inhibitory effect on the PSII quantum yield, indicating a reduction of the photosynthetic electron transport performance due to damage to the structural and functional properties of PSII (Oukarroum et al. 2015).

Likewise, Chlorella vulgaris exposed to NiO-NPs for 96 h showed cellular alterations, which were related to NiO-NP concentration (EC50 of 13.7 mg L−1). They particularly inhibited cell division (relative cell size and granularity), deteriorated the photosynthetic apparatus (chlorophyll synthesis and photochemical reactions of photosynthesis), and induced oxidative stress (ROS formation). The TEM and X-ray analysis indicated that NiO-NPs were able to cross biological membranes and accumulate inside the algal cells (Oukarroum et al. 2017). In addition, 20 nm NiO-NPs displayed severe inhibitory effect on the growth of C. vulgaris after 96 h exposure with EC50 value of 31.4 mg L−1 (Li et al. 2017). The changes observed were plasmolysis with a shriveled cell shape, disruption of plasma membrane, cytosol leakage, and disorders in thylakoid grana lamella. Moreover, NP aggregation as well as partial reduction to Ni0 could be observed, suggesting a possible remediation strategy of aquatic pollution (Li et al. 2017). These results for 20 nm NiO-NPs are comparable to the 0.35 mg NiO L−1 EC50 value obtained in a previous study for Pseudokirchneriella subcapitata (Griffitt et al. 2008). Data on toxicity effects of NiO-NPs are presented in Table 16.5.

Table 16.5 Summarized results from NiO-NP toxicity studies performed on lower plants
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3.8 Silica Oxide Nanoparticles

In order to test the hypothesis that the ecotoxicity of nanoparticles is related to their surface area and not to their mass, toxicity of silica (SiO2) nanoparticles was monitored on the growth of Pseudokirchneriella subcapitata exposed to stable silica suspensions (Van Hoecke et al. 2008). Commercial Ludox suspensions of NPs with diameters 12.5 and 27 nm were toxic, with 20% effect concentration (EC20) values on growth rate of 20.0 and 28.8 mg L−1, respectively, at 72 h. Because no aggregation was observed and the dissolution of NPs was negligible, the toxicity was attributable to the solid nanospheres. There was no significant difference in toxicity on expressing the concentration as a surface area. The 72 h EC20 values were 4.7 and 3.9 mg L−1. Silica bulk material was found to be nontoxic up to 1 g L−1. TEM studies with 100 mg SiO2-NPs L−1 (particle size 12.5 and 27 nm) elucidated no evidence of particle uptake even though the particles clearly adhered to the cell wall surface (Van Hoecke et al. 2008).

In a different study, 96 h exposure of Scenedesmus obliquus to SiO2-NPs (10–20 nm, 25–200 mg L−1) resulted in a significant concentration-dependent decrease in chlorophyll content, whereas the carotenoid content was unaffected. EC50 value could not be determined since SiO2-NPs were not toxic probably due to shading on the cell surface (Wei et al. 2010a).

4 Carbon-Based Nanoparticles

4.1 Fullerene

The effects of carbon fullerene C60 were investigated on Chlamydomonas reinhardtii. The assays included also a bioaccumulation test to observe whether the algae accumulate nanomaterials and whether they have negative effects on Daphnia magna (water flea), a small planktonic crustacean belonging to the subclass Phyllopoda and consecutively on the whole trophic chain (Luo 2007). Population changes were measured over an initial period of 48 h, which was then extended to 480 h to estimate the long-term effects. The effects were long lasting, as the algal population treated with 10 mg L−1 of C60-NPs was unable to recover within a 20-day period. C60 treatment caused color changes, cell lysis, and difficulties in reproduction. More algae died on 1 mg L−1 of C60, compared to 10 mg L−1 of C12. C60 NPs were also more toxic to D. magna than C12. Bioaccumulation studies indicate the relocation of nanomaterials from the alga to Daphnia, primarily through water but also through the alga, but the trend is not conclusive. Dynamic light scattering studies indicated aggregate formation, when the particles were introduced in an aquatic environment that led to the induction of oxidative stress.

The nanocrystalline fullerene (nC60) uptake amounts and trophic transfer efficiency to the predator (Daphnia magna) through dietary exposure to algae or algal subcellular fractions (Scenedesmus obliquus) have also been investigated (Chen et al. 2016). The nC60-contaminated algae were separated into the cell wall (CW), cell organelle (CO), and cell membrane (CM) fractions. The highest nC60 distribution was in CW, followed by CO and CM subparts. Further, the sublethal concentration for S. obliquus has been determined as 0.09 mg L−1 after 72 h exposure (Tao et al. 2015). During a sublethal experiment of C60 that was carried out for 60 days, the photosynthesis processes, the photosynthetic polysaccharide products, soluble protein, and total lipids in S. obliquus were decreased. Additionally, chlorophyll a and chlorophyll b were negatively impacted possibly due to the 40% algal Mg2+ decline at the sublethal concentration (0.09 mg L−1) of C60. The decline was due to inhibition of Mg2+-ATPase activity caused by nC60 aggregates. On the other hand, on comparing the highest and lowest sensitivity responses of a bioassay on Pseudokirchneriella subcapitata, Blaise et al. (2008) found fullerene C60 to be less toxic than other NPs and placed it in the “not toxic” category (>100 mg L−1) (Blaise et al. 2008).

Pseudokirchneriella subcapitata as well as the crustacean D. magna were used in a series of toxicity tests for studying the influence of C60 aggregates on toxicity and bioaccumulation (Baun et al. 2008). C60 powder was stirred in water over 2 months, and the aggregates formed were mixed with four environmental contaminants (atrazine, methyl parathion, pentachlorophenol-PCP, and phenanthrene) with different physicochemical properties and toxic modes of action, 5 days prior to testing. The sorption to C60 aggregates was 85% for phenanthrene and 10% for the rest of the compounds. In the presence of C60 suspensions, the toxicity of phenanthrene increased (from 720 μg L−1 to 430 μg L−1), and the toxicity of PCP decreased (from 36 μg L−1 to 70 μg L−1), and a consequent increase in toxicity was found for phenanthrene after addition of C60 to the aqueous solution. Addition of C60 suspensions reduced the toxicity of PCP. Finally, no enhanced bioaccumulation of phenanthrene was observed in the presence of C60 (Baun et al. 2008).

4.2 Carbon Nanotubes (CNTs)

In a recent study, multiwalled CNT (MWNT) material was carboxylated by microwave-assisted acid oxidations (f-MWNTs) and was examined for potential toxicity effects, using the unicellular marine chlorophyte algaDunaliella tertiolecta (Wei et al. 2010a, b). Concentrations 5 and 10 mg f-MWNTs L−1 caused substantial growth lag phase, and the EC50 value at 96 h was 0.82 mg L−1. This impact is in line with the 72 h IC25 value (1.04 mg L−1) of single-walled carbon nanotubes (SWCNTs) on the growth of green alga Pseudokirchneriella subcapitata (Blaise et al. 2008). Especially, at 10 mg L−1 f-MWNTs, 36% reduction in exponential growth rate was observed indicating the presence of oxidative stress and 22% reduction in photosystem II (PSII) quantum yield (Wei et al. 2010a, b). The results differed in a later study, where oxidized SWCNTs (f-SWCNTs) caused 30% growth inhibition, 18% decrease in the photosynthetic yield, and 95% reduction in the intracellular glutathione levels of Dunaliella tertiolecta (Thakkar et al. 2016).

In a study of growth inhibition and photosynthetic activity in Chlorella vulgaris and P. subcapitata, EC50 values were 1.8 mg CNTs L−1 and 20 mg CNTs L−1, respectively, in well-dispersed suspension whereas 24 mg CNTs L−1 and 36 mg CNTs L−1, respectively, in agglomerated suspension (Schwab et al. 2011). The photosynthetic activity was not affected, whereas growth inhibition was correlated to the shading of CNTs and the agglomeration of algal cells, suggesting that the growth might be affected by the shading caused by the CNTs and by alga-CNT agglomerates. However, in another study the toxicological effects of MWCNTs were not dose-dependent (Pereira et al. 2014). Exposure of C. vulgaris to MWCNTs induced SOD activity, decreased intracellular ATP levels, and further induced ultrastructural cell damage. Uptake of MWCNTs was observed when cells were cultured in BB medium, but this internalization was not repeated when cells were cultured in Seine river water. The toxicity of MWCNTs was also investigated in Chlorella sp. focusing on the four possible mechanisms for the algal growth inhibition (i.e., oxidative stress, agglomeration, physical interactions, and shading effects) and their correlation to the MWCNT size and concentration. At MWCNT concentrations near EC50 at 96 h, the oxidative stress accounted for approximately 50% of the algal growth inhibition, whereas 25% of it owes to agglomeration-physical interactions and 25% to the shading effects (Long et al. 2012). Moreover, toxicity of MWCNTs toward Chlorella pyrenoidosa was investigated in the presence of different dissolved organic matters, i.e., a natural originated humic acid (HA) and two synthetic surfactants [sodium dodecylbenzenesulfonate (SDBS) and octyl phenoxy polyethoxyethanol (TX100)] (Zhang et al. 2015). Cell internalization of MWCNTs and induction of oxidative stress were promoted by SDBS and TX100, while HA alleviated the MWCNT toxicity by limiting the cell internalization of MWCNTs and reducing the oxidative stress.

The acute aquatic toxicity of SWCNTs (∼20 um in length and 1 ∼ 1.2 nm in diameter) has been evaluated toward two freshwater microalgae (Raphidocelis subcapitata and Chlorella vulgaris) after a 72 h incubation period (Sohn et al. 2015). The SWCNTs inhibited the growth of R. subcapitata and C. vulgaris with EC50 values of 29.99 and 30.96 mg CNTs L−1, respectively, and were classified as “acute category 3” in the Globally Harmonized System (GHS) of classification and labeling of chemicals. A study of the effects of SWCNTs was undertaken on a population of microalga Chromochloris zofingiensis with a special focus on the profile and production of pigments and fatty acids (Wang and Yang 2013). The alga after a 6-day incubation with SWCNTs showed biomass enhancement at low concentrations (40–160 mg L−1) and inhibition at high concentrations (320 mg L−1). By contrast, fatty acids and pigments accumulation decreased over the range of the tested concentrations indicating an increasing sensitivity of the inhibitive toxicity markers as follows: biomass and fatty acids < primary carotenoids < chlorophylls < secondary carotenoids. The data recorded for toxicity of CNTs are shown in Table 16.6.

Table 16.6 Summarized results from CNT toxicity studies performed on algae
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5 Quantum Dots

Potential toxicity of quantum dots (QDs) was assessed by using Chlamydomonas reinhardtii as a model system (Wang et al. 2008). The response of the organism to QDs was initially assessed by growth kinetics that showed growth inhibition and formation of aggregates during the first 2–3 days of cultivation (EC50 value = 5 mg QDs L−1), followed by a rapid recovery and reduction of cell aggregation as the culture proceeded. Cellular oxidative stress occurred 6 h after exposure to QDs, as confirmed by the transcriptional expression profiling of three stress response genes (SOD1, GPX, and CAT). The expression of these genes was temporarily enhanced in cultures containing 0.1 mg QDs L−1, with the maximum transcripts of SOD1, GPX, and CAT occurring after 3 h of treatment, proportionally to the initial concentration of QDs. As the cultures continued, recovery in growth was observed, and the extent of recovery, as indicated by the final cell concentration, was dosage-dependent. In another study, the adsorption of carboxyl-functionalized polymer-coated QDs (CdSe/ZnS-QDs) and their effects on C. reinhardtii photosynthesis were examined (Lin et al. 2009). The amount of QDs adsorbed onto algae logarithmically depends upon the equilibrium concentration of the QDs with Freundlich constants determined as k = 0.588 ppm1−n and n = 0.629. Furthermore, CO2 depletion and O2 production assays showed a significantly inhibited photosynthetic activity of the alga exposed to QDs in concentrations above 100 ppm and 5 ppm, respectively, suggesting the potential impact of NP adsorption on the obstruction of gas flow and nutrients uptake for the algae.

In addition, when C. reinhardtii was exposed to increasing QD concentrations, dissolution increased with decreasing pH (Domingos et al. 2011). QDs were accumulated by the algal cells (in a size-dependent manner), though the particles may have been dissolved upon entry into the cells. Whole transcriptome screening using RNA-Seq analysis identified 174 transcripts that were specifically upregulated by QDs. Moreover, pathways linked to transmembrane activity, proteolysis involving proteasome activation, and ubiquitin-mediated processes were observed. In a different study, effects of CdSe/ZnS-QDs were studied on the availability of Cu and Pb on two strains of C. reinhardtii (a wall-less and a walled strain containing glycoproteins) and a microalga (Chlorella kesslerii) that possesses a cellulosic cell wall (Worms et al. 2012). The results indicated that QDs decreased the intracellular Cu and Pb contents (non-extractable by EDTA) to almost half in C. kesslerii and in the walled strain of C. reinhardtii but increased them about 3.5–4 times in the wall-less strain, suggesting that CdSe/ZnS-QDs could influence metal bioavailability due to the interactions of QDs with the cell wall.

As the stability of NPs in seawater is an important requisite for efficient interactions with living organisms, the effects of water-soluble CdSe-QDs were assessed on the marine microalga Phaeodactylum tricornutum (Morelli et al. 2012). High QD concentrations (>0.5 nM) caused a dose-dependent inhibition of growth rate and induced ROS formation as well as modulation of SOD and CAT activities. Similarly, functionalized CdSe/ZnS-QDs (amine- and carboxyl-) showed limited toxicity to the marine diatom Thalassiosira pseudonana after a 5-day exposure under varied nutrient conditions (enriched versus nitrogen-limited media) (Zhang et al. 2013). Production of proteins in T. pseudonana was induced suggesting that these extracellular proteins might be involved in the detoxification of QDs by this alga via the Cd release of QDs.

The bioaccumulation kinetics of thioglycolic acid-stabilized CdTe quantum dots (TGA-CdTe-QDs) was investigated in a freshwater alga Ochromonas danica (Wang et al. 2013). Flow cytometry measurements showed high photoluminescent intensity in cells during the exposure time (1, 5, 10, 15, 20, 30, 40, 50, 60 min) suggesting internalization of TGA-CdTe-QDs, while a significant NP uptake was observed through the mechanism of micropinocytosis. The intracellular TGA-CdTe-QDs had negligibly direct acute effects on the algae, and their toxicity was mainly caused by Cd ion liberation into the bulk medium. Quick elimination in the photoluminescent intensity of cellular TGA-CdTe-QDs was also observed, and it was correlated to QD dissolution, surface modification, or expulsion out of the cells. In another report, cytotoxicity of two types of QDs, i.e., carbon QDs (N,S-doped CQDs, N-doped CQDs, no-doped CQDs) and metal QDs (CdTe-QDs, CdS-QDs, CuInS2/ZnS-QDs), was investigated on Chlorella pyrenoidosa (Xiao et al. 2016). On treating C. Pyrenoidosa with various concentrations of QDs, the total protein and chlorophyll a contents were reduced in a dose-response manner. The EC50 values (mg L−1) of CQDs and MQDs (shown in Table 16.7) were determined by a growth inhibition biotest (algal cells counting) for 96 h, and their toxicity order was CuInS2/ZnS-QDs < no-doped CQDs < N-doped CQDs < N,S-doped CQDs < CdS-QDs < CdTe-QDs. QDs enhanced the activity of antioxidant enzyme superoxide dismutase (SOD) and decreased the reduced glutathione (GSH) level in a dose-dependent manner. Additionally, QDs enhanced the accumulation of malondialdehyde (MDA). Finally, the toxicity of CQDs was smaller than MQDs, with the toxicity of CuInS2/ZnS-QDs being the smallest one (Xiao et al. 2016).

Table 16.7 Summarized results from QD toxicity studies performed on algae
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6 Conclusions

The quick growth of nanotechnology over the years has led to rapid development of its commercial applications, which involves the use of a great variety of manufactured NPs. The use of these organic and inorganic nanomaterials may result in the surreptitious discharge of these materials into the environment through soil, sediment, and biosolids from wastewater treatment. Algae and other nonvascular plants constitute an important component of our ecosystem, and toxic effects of nanoparticles on their growth attract serious concerns. Most of the studies conducted on NP toxicity to nonvascular plants have been focused on algae; only a few could encompass bryophytes. The toxic action of NPs can involve some distinct mechanisms, but drawing a general conclusion regarding factors that determine the toxicological effects of NPs is not possible, because toxicity data generated thus far are conflicting and inconsistent. There are indications that NPs might interact directly with algae due to secondary particle size and/or specific surface area or indirectly through release of toxic substances into the exposure media. Further, the duration of exposure to NPs may be an important parameter for the assessment of their toxicity potential (even at low concentrations), which may be more representative of real environmental conditions. Additional comprehensive investigations are urgently and immensely required to examine the impact of NPs on the food chain and the environment and also to be able to reach at logical conclusions for establishing regulations over the use, confinement, and disposal of NPs for the protection of the ecosystem and humankind.