Over the last two decades, fungicide use has increased across the globe. For instance, fungicide usage for total acreage in the United States increased from around 2% in 2005 to ~ 30% by 2009 (Hooser et al., 2012). This spike in usage is due to the rise and spread of fungal pathogens that infect many crops including wheat, soybeans, and grapes. Globally, Europe and North America account for the majority of fungicide use (Zubrod et al., 2019). In Europe, a large portion of fungicide usage is for viticulture, wine-growing operations, while in North America the majority of fungicides are applied to corn, wheat, and soybeans (Adams & Bruhl, 2020; Hartman et al., 2014). Despite this rapid rise in fungicide usage, research examining their ecotoxicity is still limited. A review of pesticide-centered ecotoxicology research found that from 1991–2013 fungicide research accounted for only ~ 13% of all studies (Köhler & Triebskorn, 2013). Given the global rise in fungal pathogens and the increasing implementation of fungicides, more work needs to be done to bridge this divide and assess the environmental risk associated with fungicides (Fisher et al., 2012, 2020).
Environmental detection of fungicides in water bodies is common despite their reported short residence time. Studies of European waters have found that fungicides are detected at higher concentrations than either herbicides or insecticides (0.96 µg/L compared to 0.063 and 0.034 µg/L) (Stehle & Schulz, 2015). Post-application surface water run-off from fields has been found to exceed recommended environmental risk limits set by the United States Environmental Protection Agency (USEPA) for the fungicide pyraclostrobin (Aamlid et al., 2021). A European study examining the detection and concentration of pesticides in sediment found that the fungicides fenpropimorph and propiconazole were the most commonly detected pesticides (Kronvang et al., 2003). In surveys of major wetlands, the United States Geological Survey found several different classes of fungicides present at detectable levels with two fungicides, azoxystrobin and boscalid, being the most commonly found pesticides (Smalling et al., 2010). Furthermore, fungicides were commonly detected in the sediment of these wetlands with pyraclostrobin present at all sites (Smalling et al., 2010). In that same study, pyraclostrobin was detected in 53% of all samples collected from the water. Collectively, this research on the widespread environmental occurrence of fungicides underscores the need for studies that assess their effects on wildlife.
To date, fungicide toxicity research has predominantly focused on strobilurins (e.g., pyraclostrobin, azoxystrobin) and chloronitriles (e.g., chlorothalonil), which are the most commonly used chemicals. Pyraclostrobin is highly toxic to several commonly tested species (Zebrafish, Danio reio; Zooplankton, Daphnia magna; Fathead Minnows, Pimephales promelas). For instance, experiments with D. magna showed a lethal concentration (LC50) of 14 µg/L in 96-hour tests (Ochoa-Acuña et al., 2009). Consistent with these results, experiments on other commonly used model organisms, such as zebrafish, show similar levels of acute toxicity (Li et al., 2018; Zhang et al., 2020). Additionally, the pyraclostrobin formulation Headline was highly toxic in amphibian larvae with 96-hour LC50 tests ranging from 2.1–15 µg/L (Belden et al., 2010; Cusaac et al., 2016; Hooser et al., 2012). Chlorothalonil studies have shown more varied results, but overall, the chemical is highly toxic with acute effects ranging from 25.5–396 µg/L (Key et al., 2003; Méndez et al., 2016). Studies with a commercial formulation of chlorothalonil (Bravo 500) have shown a range of values for different species (Blue mussel, Mytilus edulis = 35 mg/L; Daphnia magna = 130–200 µg/L) (Ernst et al., 1991). Further chlorothalonil studies with amphibians also reported LC50 values ranging from 59–340 µg/L (Acquaroni et al., 2021; Ghose et al., 2014) for commercial formulations and 26–32 µg/L using pure chlorothalonil (Méndez et al., 2016). While more work is needed to generalize toxicity values among groups, fungicides are generally more toxic than other pesticide groups with LC50 values measured in µg/L compared to mg/L. It is also important to note that these low LC50s are within the range of the expected environmental concentration for many of these fungicides (Deb et al., 2010; McMahon et al., 2011).
Pyraclostrobin and chlorothalonil are two of the most recommended fungicides for the treatment and prevention of different forms of rust affecting multiple crops and are widely used across the United States (Deb et al., 2010; Hartman et al., 2014; Willming & Maul, 2016; Zubrod et al., 2019). Thus, we conducted acute toxicity tests on larvae of six amphibian species using these two fungicides. To date, only eight anuran species have been included in acute studies involving these fungicides, and the results are varied (Table 1). Moreover, no salamanders have been tested. While the existing data suggest that fungicides are acutely toxic to amphibians, the inclusion of more species will help generate generality. Moreover, there is limited research on species from the midwestern United States where fungicides usage is widespread in agricultural operations (Deb et al., 2010; Hartman et al., 2014; Hooser et al., 2012). Testing species that live in habitats that are most likely to be impacted by these fungicides will help inform risk assessment. Finally, previous work generally has focused on the commercial formulations of both chlorothalonil and pyraclostrobin. Because commercial formulations contain many different types of additives that can alter the toxicity of the mixture or be toxic substances alone, research conducted with technical grade compounds can directly assess fungicide toxicity independent of the additives.
Table 1
Pyraclostrobin and chlorothalonil toxicity study resultant LC50s
Species | Family | Pesticide | LC50 (µg/L) | Reference |
Bufo cognatus | Bufonidae | Pyraclostrobin | 3.7–15 | Belden et al. 2010 |
Bufo cognatus | Bufonidae | Pyraclostrobin | 10 | Hooser et al. 2012 |
Acris blanchardi | Hylidae | Pyraclostrobin | 2.1 | Cusaac et al.2016 |
Agalychnis callidryas | Phyllomedusidae | Chlorothalonil | 59.36 | Méndez et al. 2014 |
Agalychnis callidryas | Phyllomedusidae | Chlorothalonil | 26.6 | Méndez et al. 2016 |
Isthmohyla pseudopuma | Hylidae | Chlorothalonil | 25.5 | Méndez et al. 2016 |
Smilisca baduinii | Hylidae | Chlorothalonil | 32.3 | Méndez et al. 2016 |
Xenopus laevis | Pipidae | Chlorothalonil | 8.2–14.4 | Yu et al. 2013 |
Spea mutliplicata | Scaphiopodidae | Chlorothalonil | 10.7 | Yu et al. 2013 |
Rhinella arenarum | Bufonidae | Chlorothalonil | 340 | Acquaroni et al. 2021 |