Differentiating pollutant-induced effects from non-contaminant stress responses in aquatic midges (Diptera: Chironomidae)

Submitted by Bryant Samuel Gagliardi

Submitted in fulfillment of the degree of Doctor of Philosophy

June 2017

School of BioSciences Faculty of Science The University of Melbourne

Abstract

Ecotoxicology is the study of environmental contaminants and their effects on organisms. Laboratory ecotoxicology typically investigates contaminant responses in standard, model test organisms. There is often a lack of understanding of the role of non-contaminant factors in inducing these stress effects. In the freshwater model insect Chironomus (Diptera: Chironomidae), deformed larvae have been observed since the early 1970s, and these have been proposed as a potentially useful sublethal contaminant response for monitoring studies. However, despite decades of research, the causal links between contaminants and deformities are uncertain, limiting the application of this endpoint in ecotoxicology. There is also a poor understanding of the role of non-contaminant environmental stressors in inducing deformities. Additionally, inbreeding depression has been shown to be a widespread compromising factor in Chironomus assays investigating lethality, reproduction and deformity effects. Inbreeding in cultures appears to be a difficult issue to resolve.

The aims of this thesis were to clarify the causes of larval deformities in Chironomus, considering chemical and non-chemical causal factors, and to investigate a potential solution to the widespread issues surrounding inbreeding in chironomid ecotoxicology. Firstly, I conducted a meta-analysis of published laboratory Chironomus deformity assays, characterising and quantifying the inconsistency of published data across > 45 years of research. I also aimed to diagnose potential causes of this inconsistency between assays, focussing on the influence of potential extraneous laboratory stressors in assays, and mortality of larvae in assays as a potential confounding variable. Significant inconsistency was observed across published assays, and mortality effects and extraneous stressors were found to be potentially contributing to this inconsistency.

Secondly, I conducted laboratory assays investigating chemical and non-chemical deformity causes in the Australian model chironomid C. tepperi. I aimed to address the key issues identified in the meta-analysis as limiting to the understanding of deformity causes. These issues were the presence of extraneous laboratory stressors (as indicated by high control deformity incidences), potential confounding

Page 1 of sublethal deformity assays by mortality (avoided by ensuring sublethal conditions), and the widespread failure to consider non-contaminant causes (addressed by testing non-chemical factors: food limitation and onset of pupation). None of the tested factors observed to induce significant deformity effects.

Finally, I aimed to investigate a potential solution to prevalent inbreeding issues in ecotoxicology associated with the use of Chironomus as a model insect. This involved developing a standard culturing methodology, and conducting preliminary toxicological investigations for the cosmopolitan nonmodel chironomid Paratanytarsus grimmii. This species reproduces asexually, by apomictic parthenogenesis. No mating takes place, meaning inbreeding is therefore avoided, suggesting P. grimmii as a candidate international model species that will enable the investigation of contaminant effects in the absence of inbreeding. A replicable P. grimmii culturing methodology was developed, and basic toxicological assays were conducted.

Page 2 Declaration

This is to certify that: i. the thesis comprises only my original work towards the PhD except where indicated in the preface, ii. due acknowledgement has been made in the text to all other material used, iii. the thesis is fewer than 100 000 words in length, exclusive of tables, maps, bibliographies and appendices.

Bryant Gagliardi

June 2017

Page 3 Preface

This thesis comprises two published research papers and one submitted research paper as results chapters; and one published discussion paper—which incorporates components of the thesis Introduction—in the supplementary information.

The published research papers are: Bryant S. Gagliardi, Vincent J. Pettigrove, Sara M. Long, & Ary A. Hoffmann (2016). A meta-analysis evaluating the relationship between aquatic contaminants and chironomid larval deformities in laboratory studies. Environmental Science & Technology, 50(23), 12903-12911. (Chapter 2)

Bryant S. Gagliardi, Sara M. Long, Vincent J. Pettigrove, & Ary A. Hoffmann (2015). The parthenogenetic cosmopolitan chironomid, Paratanytarsus grimmii, as a new standard test species for ecotoxicology: Culturing methodology and sensitivity to aqueous pollutants. Bulletin of Environmental Contamination and Toxicology, 95(3), 350-356. (Chapter 4)

The published discussion paper is: Bryant S. Gagliardi, Ary A. Hoffmann, & Vincent J. Pettigrove (2016). Inbreeding depression as a compromising factor in ecotoxicological assays. Integrated Environmental Assessment and Management, 12(3), 595-597. (Appendix 1)

The submitted research paper is: Bryant S. Gagliardi, Sara M. Long, Vincent J. Pettigrove, & Ary A. Hoffmann (in review). A re-evaluation of chironomid deformities as an environmental stress response: Sublethal toxicological assays and noncontaminant factors. Environmental Pollution (Chapter 3)

The content of these papers is my own work and co-authorship represents a supervisory role of the other authors. Other authors provided scientific advice, trained me in laboratory methods, assisted with statistical analyses, and reviewed the manuscript before submission.

Page 4

Acknowledgements

Firstly, my immense gratitude goes to my supervisors Ary Hoffmann, Sara Long and Vin Pettigrove. Your guidance and mentorship, not to mention patience, has been instrumental in seeing me through the PhD. Your expertise and dedication to your respective fields of science has served as an ongoing inspiration to me, and will continue to do so well beyond the submission of this thesis. Thanks also to my advisory committee members Mick Keough and Kath Hassell.

Thank you to my colleagues at CAPIM, who helped in a million ways great and small, throughout this project. So many times the seemingly smallest piece of advice saved hours of wasted effort, or got me around an apparently dead-end in my research. Thanks especially to Bec Reid, Tyler Mehler, Steve Marshall, Simon Sharp, Kallie Townsend and Mel Carew for such advice. Big thanks also to Lisa Golding, Katy Jeppe, Pat Bonney, and Bec Reid for collecting the original wild P. grimmii stocks, and to Georgia Sinclair for assistance in the field. I’m also grateful to Emily Thomson and Nick Bell for much administrative help. Your helpfulness can’t really be overstated, as the red tape side of things, to put it mildly…can sometimes frustrate me. Gratitude to Teresa Norberg-King at US EPA Duluth, Minnesota for the tour of facilities, and imparting much knowledge on chironomid ecotoxicology.

Thanks to my loving family, especially Mum and Dad who fostered an interest in biology from a young age, be it through purchasing my first microscope as a kid, or taping all those David Attenborough documentaries. And, not least of all, thank you to my wife Pip Griffin, who played the invaluable dual roles of emotional supporter and academic advisor (not to mention, multi-panel R-plot layout arranging assistant). I doubt I could have gotten through the previous four years without you, here’s to many more outside the shadow of the PhD!

This PhD was funded by the Melbourne Research Scholarship and the Centre for Aquatic Pollution Identification and Management, with additional funding provided by the Holsworth Wildlife Research Endowment, the University of Melbourne Faculty of

Page 5 Science Travel Abroad Scholarship, the Pest and Environmental Adaptation Research Group and the Society of Environmental Toxicology and Chemistry Australasia Student Travel Award.

Page 6 Contents

Abstract……………………………………………………………………………………...1 Declaration...... 3 Preface……………………………………………………………………………………….4 Acknowledgements………………………………………………………………………..5 Contents……………………………………………………………………………………..7 Chapter 1: Introduction………………………………………..……………………….....8 1.1 Freshwater ecotoxicology………………………………………………………8 1.2 Chironomidae in freshwater ecotoxicology…………………………………10 1.3 Chironomid deformities as an ecotoxicological endpoint………………….11 1.4 Inbreeding as a compromising factor in ecotoxicological assays………...15 1.5 Paratanytarsus grimmii as a potential test chironomid…………………….17 1.6 Thesis aims and overview…………………………………………………….20 1.7 References……………………………………………………………………..21 Chapter 2: A meta-analysis evaluating the relationship between aquatic contaminants and chironomid larval deformities in laboratory assays………..31 Chapter 3: A re-evaluation of chironomid deformities as an environmental stress response…..………………………………………………………………………41 3.1 Abstract…………………………………………………………………………42 3.2 Keywords……………………………………………………………………….42 3.3 Introduction……………………………………………………………………..43 3.4 Materials and Methods………………………………………………………..46 3.5 Results………………………………………………………………………….53 3.6 Discussion……………………………………………………………………...54 3.7 Acknowledgements……………………………………………………………59 3.8 References……………………………………………………………………..60 3.9 Tables…………………………………………………………………………..66 3.10 Figures………………………………………………………………………...67

Chapter 4: The parthenogenetic cosmopolitan chironomid, Paratanytarsus grimmii, as a new standard test species for ecotoxicology………………………71

Chapter 5: General discussion………………………………………………………...79 5.2 References……………………………………………………………………..89

Appendix 1: Discussion - Inbreeding depression in ecotoxicology…………….93 Appendix 2: Supplementary material to Chapter 2…………………………………97 Appendix 3: Supplementary material to Chapter 3……………………………….108

Page 7 Chapter 1 - Introduction

Freshwater ecotoxicology Anthropogenic activities threaten global biodiversity (Tittensor et al., 2014; Newbold et al., 2015). Pollution—the release of toxic contaminants into the environment—is a key anthropogenic threat to ecosystems worldwide (Azevedo et al., 2013; Janse et al., 2015; Stehle and Schulz, 2015). Polluted environments are consistently species- poor relative to unpolluted environments (Ficken and Byrne, 2013; Zhang et al., 2015), and pollution remediation efforts often result in increased species richness relative to the contaminated state (Adams et al. 2005; Stockdale et al., 2014; Rose et al., 2016), demonstrating pollution as an important driver of biodiversity loss.

Freshwater habitats (rivers, lakes and wetlands) harbour significant biodiversity, containing 6 % of the earth’s species, despite covering only 0.8 % of its surface area [reviewed by (Dudgeon et al., 2006)]. These habitats are particularly vulnerable to pollution effects. Sixty-five percent of earth’s continental river discharges, and the aquatic habitats supported by this water, are under threat from anthropogenic activities including pollution (Vörösmarty et al., 2010). Freshwater habitats are vulnerable as they often occupy low-lying areas. This sees them receiving contaminants from not only nearby sources, but from the entire surrounding landscape (Allan, 2004; Dudgeon, 2014) via point discharges, surface runoff, aerial drift, atmospheric deposition and groundwater inputs (Kookana et al., 1998). In addition, freshwater habitats typically lack the large water volume of marine habitats, meaning they have less capacity to dilute contaminants (Dudgeon, 2014). Perhaps due to these factors, freshwater species appear to suffer a disproportionate extinction rate, far exceeding that observed in terrestrial taxa (Ricciardi and Rasmussen, 1999; Dudgeon et al., 2006).

Addressing the effects of aquatic pollution requires the accurate monitoring and quantification of its effects on biota. The health of aquatic ecosystems—in relation to contamination—in some regions is routinely monitored by government or management bodies. Most often, this “biomonitoring” involves surveys of macroinvertebrate communities (Rosenberg and Resh, 1993; Buss et al., 2014),

Page 8 assessing presence and absence of macroinvertebrate families. This approach can suggest pollution as causal to observed alterations in community composition or structure (Bunzel et al., 2013). However, effects observed at the community level of ecological organisation often lack a mechanistic understanding of the casual factors involved (Clements, 2000), and aquatic systems are typically influenced by a host of non-contaminant environmental stressors (e.g. altered flow regimes, introduced species, erosion and habitat degradation, climate change impacts etc.) in addition to pollution. As a result, non-contaminant stressors cannot always be eliminated as causal to observed biodiversity changes (Godfrey, 1978; Chadwick and Canton, 1984), limiting our understanding of the specific role played by pollution in aquatic ecosystem degradation. This in turn makes it difficult for environmental managers to effectively direct (often limited) financial resources to addressing the most problematic environmental issues faced in a given aquatic system.

Ecotoxicology is the study of contaminants and their effects on organisms (Newman and Evans, 2006). To better understand the threats posed by contaminants, much work in this discipline is dedicated to investigating cause-effect relationships between contaminants and organism responses in the laboratory (Banks and Stark, 1998; O'Brien and Keough, 2014). Chemical causality is simpler to determine in individual-level studies, as the contaminant-individual relationship is inherently less complex than the contaminant-community relationship (which involves many indirect effects) (Clements, 2000; Koperski, 2011). Individual-level laboratory assays also have the advantage of excluding environmental variables that may complicate or confound results in field-based experiments (Chapman, 1995). Laboratory experiments have hence proven useful for demonstrating the lethal and sublethal effects of pollutants to a wide variety of test organisms at the individual level (Maltby, 1999).

Lower-level (subcellular and individual-level) studies provide powerful tools determining causality in ecotoxicology, and a “starting point” for demonstrating contaminant-induced ecological effects. The most severe pollution response (or, “endpoint”) is mortality. However, there are also many sublethal responses exhibited by organisms, which may contribute to more insidious, chronic declines in individual, population or community health. For example, in terrestrial surveys, birds exposed to

Page 9 organochlorine pesticides (e.g. DDT) in agricultural environments in the 20th century were found to lay eggs with thinner shells, leading to reduced reproductive success and severe population declines (Ames, 1966; Porter and Wiemeyer, 1969). The overall abundance of such ecotoxicological data—clearly linking individual-level effects (e.g. laying of thinly shelled eggs), to “ecologically relevant” higher-level effects (e.g. population declines)—is probably unsatisfactorily low (Köhler and Triebskorn, 2013). Nonetheless, ecotoxicological studies have clearly demonstrated that contaminants can induce physiological (Heath, 1995), behavioural (Hellou, 2011) and genetic (Li et al., 2016) changes in organisms which have potentially deleterious consequences for the individual, and may also impact upon population, community and ecosystem health (Calow and Sibly, 1990).

Chironomidae in freshwater ecotoxicology Chironomids, or “non-biting midges” (Diptera: Chironomidae), are a species-rich aquatic insect family (Armitage et al., 1995), whose species vary in their tolerance to pollution (Pettigrove and Hoffmann, 2005). They inhabit almost every type of freshwater habitat on earth (Armitage et al., 1995), and are ecologically important, particularly as prey for many vertebrate species (Tokeshi, 1995). They are therefore commonly used as test organisms in laboratory and field-based aquatic ecotoxicological experiments (Lindegaard, 1995). Their juvenile (egg, larval and pupal) life stages are aquatic and therefore exposed to any pollutants in the water column (Pinder, 1986; Pinder, 1995). Many species inhabit the benthos (Pinder, 1986), and hence are also commonly used in sediment ecotoxicology studies (OECD, 2004; Pettigrove and Hoffmann, 2005).

Chironomus is the international “standard” laboratory ecotoxicological test chironomid genus, in that there are standardised methodologies published by bodies such as the Organisation for Economic Co-operation and Development (OECD, 2004); the US EPA (US EPA, 2000); and the American Society for Testing and Materials International (ASTM, 2007). Chironomus are typically abundant in urban and agricultural freshwater habitats (Casper, 1994; Stevens et al., 2006), relatively large and conspicuous as larvae, and robust to laboratory conditions (Taenzler et al., 2007). This makes them relatively easy to collect from the wild and amenable to

Page 10 laboratory studies. As such, the vast majority of chironomid ecotoxicological studies use Chironomus spp. (Lindegaard, 1995).

The endpoints prescribed for assessment in standard Chironomus methodologies are mortality, immobilisation, emergence (from pupal to adult stage) and growth. These endpoints are commonly utilised in applied and academic studies. Increasingly in chironomid (as with most other areas of) ecotoxicology, there is growing effort towards determining novel, sublethal contaminant responses, such as behavioural effects (Bisthoven et al., 2004), gene expression differences (Park and Choi, 2017), biochemical metabolic effects (Long et al., 2015) and morphological changes (Pracheil et al., 2016). This effort is driven by the need to develop meaningful assays, that are reliably sensitive to contaminant exposure (at environmentally-relevant concentrations), that give a meaningful indication of organismal stress, and that are specific enough to differentiate effects of different contaminant types in mixed exposure scenarios (Hines et al., 2010; Jeppe et al., 2014). Sublethal endpoints are thought to provide an “early-warning” of more dramatic mortality/population extinction events. In a management context, they may therefore also enable the early diagnosis of environmental problems, before they become more difficult—and expensive—to remediate (Hellou, 2011).

Chironomid deformities as an ecotoxicological endpoint Morphological deformities in chironomid larvae have been widely investigated as an ecotoxicological endpoint (Janssens de Bisthoven and Gerhardt, 2005). This work has primarily investigated metal and pesticide effects in Chironomus. Since the early 1970s, abnormalities [of the mouthparts and/or antennae, see Figure 1 in Vermeulen (1995)] have been observed in field-collected chironomid larvae (Hamilton and Saether, 1971). There is often a spatial association between contaminants and deformity incidence in the field (Janssens de Bisthoven et al., 1995), suggesting a causal relationship, and hence that deformities may be a useful ecotoxicological endpoint. Ecotoxicological investigations of chironomid deformities has consequently been an area of considerable research effort. A literature search of the databases Google Scholar, Web of Science, and Scopus, using the search terms “chironom* + deform*” and “chironom* + abnorm*” reveals 142 peer-reviewed articles and text

Page 11 book chapters—investigating chironomid larval deformities as a response to contaminant exposure—published prior to 1 January 2016.

It has been suggested that chironomid deformities possess many characteristics that make them desirable as a novel ecotoxicological endpoint, for both laboratory and field assessments. These suggested benefits include contaminant sensitivity (Warwick, 1985), their potential indication of organismal stress (Janssens de Bisthoven et al., 2001), and the potential to differentiate effects of different contaminant types in mixed exposure scenarios (Vermeulen, 1995). Additionally, screening of deformities in field chironomid populations, and assessing for differences in deformity frequency between contaminated and uncontaminated sites, is a relatively inexpensive pollution screening method. When compared with chemical, ecological, microbial, genetic and biochemical assessments, deformity assessments require relatively less sophisticated equipment and taxonomic expertise. For this reason, chironomid deformities have been proposed as a viable pollution monitoring method in developing nations (Janssens de Bisthoven, 1999). Another advantage of chironomid deformity field assessments is the near-ubiquity of chironomids in freshwater habitats of all types (Armitage, 1995), meaning resident populations can be easily assessed, without the necessary requirement for culturing test organisms in elaborate laboratory systems.

Despite their promise as an environmental monitoring tool, and over four decades of research, the utility of chironomid deformities in ecotoxicology remains in question. This is primarily due to an apparent inconsistency in experimental results. Field results variously suggest positive and negative associations between contaminant concentrations and deformity incidence (Janssens de Bisthoven et al., 1995), rather than consistently indicating positive dose-response effects. To investigate effects in the absence of potential confounding environmental factors, laboratory experiments have been conducted since the first observations of deformities (Hamilton and Saether, 1971). However, the apparent inconsistency of field results has also been widely observed in laboratory work, and data do not routinely support contaminant causality. These complex challenges appear to be echoed in investigations of contaminant-induced deformities in amphibians (Blaustein and Johnson, 2003). By

Page 12 comparison, fish deformities appear to be relatively well-established as indicators of contaminant exposure (Sfakianakis et al., 2015).

Chironomid deformity data are inconsistent in that, while some studies indicate positive associations between contaminant concentration and deformity frequencies in laboratory or field populations, others indicate non-significant, or even negative associations. For example, Janssens de Bisthoven et al. (1995) observed a positive association between DDE (toxic breakdown product of DDT) concentration and deformity frequencies in the field, but Warwick (1985) observed a negative association in the laboratory. Similarly inconsistent results have been observed in laboratory copper (Vermeulen, 1995) and mixed contaminant exposures (Martinez et al., 2004). These results bring into question the reliability with which a high frequency of deformities indicates population exposure to contaminants. For this reason, the consensus view among researchers is that chironomid deformity data are too inconsistent to be routinely applied as an ecotoxicological endpoint (e.g. Hämäläinen, 1999; Janssens de Bisthoven et al., 2001; Arambourou et al., 2012), despite the proposed benefits of this method.

Although the inconsistency of deformity data is the subject of much discussion, the extent of this inconsistency—and its potential causes—have rarely been empirically investigated. For example, while the occurrence of negative/non-significant associations is often raised as evidence of inconsistency, it is not known whether these types of results are simply non-representative “outliers”, or whether they are sufficiently common in the literature as to bring contaminant causality into question. Similarly, the reasons why there is inconsistency across results have mostly been speculated upon, rather than tested.

Differences between human observers in their designation of larvae as either “deformed” or “nondeformed” has recently been demonstrated as one potential source of inconsistency (Salmelin et al., 2015). Other factors that have been hypothesised include potentially differing water quality conditions (temperature, pH, dissolved oxygen concentration, salinity levels) between experiments (Vermeulen, 1995); and other stressors impacting laboratory populations such as malnutrition, inbreeding depression, parental effects and unknown stressors in substrates or food

Page 13 (Janssens de Bisthoven and Gerhardt, 2005). There is some experimental data to support inbreeding depression (Vogt et al., 2013) and parental effects (Servia et al., 2000) as influences on chironomid deformity results, however a thorough investigation of these factors and their influence upon chironomid deformity results remains to be conducted.

In addition to background stressors, mortality of larvae in experiments is a potentially confounding factor, that has been considered even more rarely than background stressors. Deformities are a sublethal endpoint, and as such should be assessed under sublethal conditions (Abdel-Moneim et al., 2015). This involves using lower than lethal contaminant concentrations, and ensuring laboratory conditions are sufficiently unstressful as to not induce mortality. The importance of not inducing mortality in sublethal assays appears to be largely overlooked in chironomid deformity assays, and possibly in ecotoxicology more broadly.

Mortality may confound deformity assays as it excludes more sensitive, “mortality- prone”, larvae from the analysis (as they do not reach the larval stage required for deformity assessment). These mortality-prone larvae probably have physiological differences from those that survive a given stressful condition (e.g. a given chemical concentration) (Polak et al., 2002). “Likelihood of exhibiting deformities” may be one of these differences, hence mortality-inducing assays select for “mortality-resistant” larvae, rather than assessing effects in the whole population. Such assays therefore risk false positive (if mortality-resistant larvae have a higher likelihood of exhibiting deformities) or false negative (if mortality-resistant larvae have a lower likelihood of exhibiting deformities) results. The latter (false negative) situation has been hypothesised to be the case in two studies, as a possible explanation for negative contaminant-deformity associations [uranium (Dias et al., 2008) and lead-zinc mixtures (Martinez et al., 2004)]. The “false positive” risk does not appear to have been considered in the literature, despite Janssens de Bisthoven et al. (2001) observing a clear positive association between mortality and deformity frequencies in cadmium assays (in this case in the absence of any clear cadmium-deformity association). It is therefore clear that deformity frequencies, in at least some cases, may simply be an artefact of mortality rates. In other words, if a significant deformity effect is observed in a lethality-inducing experiment, it cannot be ascertained

Page 14 whether the contaminant has mechanistically induced that effect, or simply selected for individuals carrying that deformity type. Across the literature however, the importance of avoiding lethality in deformities is not broadly considered, as studies commonly induce mortality effects (Janssens de Bisthoven, 1999), or do not quantify/report mortality levels.

These potentially compromising/confounding factors—“background stressors” and mortality—could contribute to inconsistency between results if they are occurring at substantially different levels between experiments. An understanding of their contribution to results is therefore crucial to clarify the role of contaminants in inducing chironomid larval deformities. Non-contaminant factors can have a considerable influence on chironomid toxicological assays (Hale et al., 2014), and it is imperative to characterise these effects and their bearing upon experimental outcomes. There is, for example, an advanced understanding of the role of dietary factors in inducing fish deformities (Cahu et al., 2003), which probably has contributed to the relatively clear understanding of the involvement of contaminants in inducing fish deformities. By comparison, the understanding of the effects of broader environmental factors with regards to chironomid deformities is poor, and appears to have contributed to over 45 years of scientific uncertainty surrounding toxicological work in this area.

Inbreeding as a compromising factor in chironomid assays Inbreeding is a common issue in small captive animal populations which suffer from genetic impoverishment (Willi et al., 2006), and this appears to be the case for Chironomus cultures (Nowak et al., 2007b). Inbreeding is the result of reproduction between two closely related parents. Inbred populations can suffer inbreeding depression (ID): the reduction in fitness due to an increase in genotypes that are homozygous for recessive deleterious traits (Charlesworth and Charlesworth, 1987). Inbreeding depression appears to act to compromise deformity toxicology assay results, by inducing a deformity effect (Vogt et al., 2013). Inbreeding depression may therefore be one of the “background stressors” contributing to inconsistent/unreliable chironomid deformity data.

Page 15 In addition to deformity assays, inbreeding effects can also compromise mortality, emergence time, egg clutch size, sex ratio and population growth rate Chironomus assays. This compromising effect comes about by inbreeding effects themselves inducing these endpoint responses, and/or potentially by interacting with contaminant toxicity [by increasing the sensitivity of inbred individuals to contaminants (Nowak et al., 2007a)]. For deformities, inbreeding depression appears to induce an effect independently of contaminant exposure [as demonstrated in a tributyltin study by Vogt et al. (2013)]. This means that a deformity potentially compromised by ID in test larvae should be identifiable by a high control deformity rate.

However, for some endpoints, inbreeding appears to only induce a deleterious effect in conjunction with contaminant exposure (suggesting additive or synergistic effects). For example, in cadmium assays, Nowak et al. (2007a) demonstrated that inbred and outbred larvae did not necessarily exhibit significantly different mortality rates in control (i.e. cadmium free) conditions. However, inbred larvae showed a significantly higher mortality rate at higher cadmium concentrations. This indicates that, for certain endpoints, an ID-compromised assay will not necessarily be identifiable by a high stress level in controls. For such endpoints, ID can therefore act as a “hidden” compromising factor. Furthermore, inbreeding appears difficult to manage in chironomid cultures. While inbreeding in captive animal populations can be alleviated by bolstering breeding stocks with “fresh material” (i.e. individuals either from the wild or external laboratory colonies), this approach appears insufficient for restoring Chironomus populations to satisfactory levels, i.e. those comparable with wild populations (Nowak et al., 2007b). These impoverished laboratory populations may also not be representative of the genetically diverse wild populations they are supposed to reflect in environmental studies (Brown et al., 2009).

Inbreeding in Chironomus cultures therefore appears to be a particularly challenging issue to monitor and manage. By sensitising test organisms, it risks alarmist conclusions regarding contaminant toxicity (Brown et al., 2009). By inducing high stress levels in control larvae (for some endpoints) (Nowak et al., 2007a), it may also risk reducing the observed effect size between control and exposed organisms (see Appendix 1), and therefore underestimating contaminant toxicity. However, most

Page 16 Chironomus ecotoxicology results rely upon laboratory-cultured colonies, and most published studies do not give explicit consideration to ID as problematic to the interpretation of final results. This is particularly concerning for situations where laboratory data is used to generate contaminant “trigger values” for monitoring/regulatory purposes, which when exceeded are supposed to indicate potential threats to aquatic ecosystems. Inbreeding depression may bring many published toxicology results for chironomids, along with other laboratory-reared species, into question, and no clear solution to this problem has yet been presented. For further discussion of the problems posed by ID to ecotoxicology, see Appendix 1.

Paratanytarsus grimmii as a potential standard test chironomid in ecotoxicology Inbreeding may be inevitable in cultures of the sexually-reproducing Chironomus: ten out of ten surveyed cultures from the United States and Europe exhibited genetic impoverishment, and experiments showed introduction of fresh material to be insufficient for alleviating this impoverishment (Nowak et al., 2007b). However, the use of other chironomid species may offer an alternative solution to concerns around inbreeding. Prior to the standardisation of chironomid ecotoxicology protocols (around the 2000s), a somewhat greater variety of laboratory test taxa than Chironomus were used (Anderson et al., 1980; Hatakeyama, 1987; Timmermans and Walker, 1989). These species included Tanytarsus dissimilis in the United States (Anderson et al., 1980) and Paratanytarsus parthenogenetica/parthenogeneticus in Japan (Sato and Yasuno, 1979). These two particular “different taxa” would later be demonstrated to actually be the same species, Paratanytarsus grimmii, a cosmopolitan chironomid often found to inhabit drinking water treatment plants (Langton et al., 1988).

Paratanytarsus grimmii, unlike the sexually-reproducing Chironomus, is an apomictic parthenogen (Porter, 1971). It therefore reproduces clonally: mothers produce daughters genetically identical to themselves. Reproduction involves mitosis, and there is no meiotic recombination of genetic material nor mating (Koltunow, 1993). Maintenance of the parental genome means that heterozygosity is “fixed” in subsequent generations (Saura et al., 1993), and ID is thus avoided. The use of P.

Page 17 grimmii in ecotoxicology therefore presents the opportunity to generate data for which the is no concern of confounding/compromising of results by ID.

Although a number of P. grimmii ecotoxicological studies were published during the 1970s – early 1990s (e.g. Hatakeyama and Yasuno, 1981; Poston et al., 1986), this output lessened with the emergence of Chironomus as the standard test genus. For example the US EPA, at which P. grimmii toxicological work in the U.S. was conducted, ceased P. grimmii cultures in favour of C. dilutus cultures around this time (T. Norberg-King, US EPA, personal communication). Chironomus dilutus and C. riparius are now used as standards in North America and Europe, and C. yoshimatsui is used in Japan (OECD, 2010). It was only sometime after these changes (favouring use of Chironomus, which started around 2000 e.g. [US EPA, 2000)] that revelations of widespread, problematic inbreeding issues emerged, around 2007 (Nowak et al., 2007b). The use of P. grimmii would avoid these issues but this potential has not been recognised. Part of the reason may be the confused taxonomic history of this species (it was recorded as a different species in many different regions across its wide geographical range), and its unusual status as an internationally cosmopolitan chironomid (Langton et al., 1988).

As well as being able to completely avoid inbreeding depression, P. grimmii has other features which make it attractive as an ecotoxicological model insect. Its cosmopolitan distribution [inhabiting all continents except Antarctica (Langton et al., 1988; Hamerlík et al., 2011)] for example, means that it could be used globally. Chironomus spp., conversely, are regionally endemic, and species differ in their geographic distribution, life histories and contaminant sensitivity. This means that toxicological results (Watts and Pascoe, 2000), and culturing and test methodologies (OECD, 2010), differ regionally according to endemic Chironomus test species. Regionally endemic model species also have a limited ecological relevance and availability outside of their native range (Freitas and Rocha, 2011). The cosmopolitan distribution of P. grimmii means that it is widely available, and toxicological data are potentially applicable to many freshwater systems globally (i.e. those in its native range). Additionally, P. grimmii has a naturally low genetic diversity across its range (Carew et al., 2013), meaning there are fewer concerns regarding genetic impoverishment in cultures, reducing issues around representativeness of wild

Page 18 populations. Additional advantages associated with use of clonal model organisms include the potential to enhance experimental precision (Forbes and Forbes, 1993) and to investigate epigenetic effects (Vandegehuchte et al., 2009).

Although there is some published ecotoxicological work using P. grimmii, as previously stated, much of it was published prior to the standardisation of chironomid protocols. These standard protocols have been since published in recognition of the importance of producing reliable, reproducible results (Burton et al., 1996). This involves the standardisation of test and culturing procedures (Soares and Calow, 1993; OECD, 2010). Published P. grimmii work, however, is regionally fragmented. Furthermore, published methods are not readily reproducible outside the laboratory they were conducted. For example, U.S. methods require the use of natural (Lake Superior) water as culture and test solution (Gersich et al., 1989), and Japanese methods require feeding of larvae with live cells of a specific alga (Hatakeyama and Yasuno, 1981). Use of differing culturing methodologies between laboratories can lead to non-repeatable results (Bradley et al., 1993), and the use of commercial culturing components (e.g. foods, salts for test solutions) is preferred over natural components to better enable standardisation (Soares and Calow, 1993; OECD, 2010).

There are no standardisable culturing or test methodologies available for P. grimmii that are based exclusively upon commercially available components. This is in contrast to the model genus Chironomus. It is not known whether existing Chironomus methods are suitable for P. grimmii. Furthermore, published P. grimmii toxicological work has not been conducted under conditions comparable to published Chironomus assays, meaning the contaminant sensitivity of P. grimmii cannot be properly compared to that of Chironomus. Pesticide work on this species has involved older insecticides (such as organochlorines) in experiments (Sato and Yasuno, 1979), and there is no data available for modern insecticides such as neonicotinoids. Baseline work is therefore needed before P. grimmii can be established as a model ecotoxicological chironomid. Once such baseline work has been completed, internationally applicable toxicological data might be generated with this species. The literature of laboratory aquatic ecotoxicology work primarily investigates effects in Chironomus as a model insect. In these assays, it is uncertain

Page 19 which observed effects are attributable to contaminants, which are the result of ID, and which are the result ID-contaminant interactive effects (Nowak et al., 2007a; Vogt et al., 2013). Use of P. grimmii will allow the isolation of contaminant effects and their examination in the absence of inbreeding effects, improving our understanding of invertebrate ecotoxicology.

Thesis aims and overview The aims of this thesis were to clarify the causes of Chironomus larval deformities, and to investigate a potential solution to issues surrounding inbreeding in chironomid ecotoxicology. Specifically, I aimed to address over 45 years of uncertainty regarding the links between contaminant exposure and deformities, by analysing trends in the published data and conducting carefully controlled laboratory experiments. To address inbreeding issues in ecotoxicology, I investigated the potential for the asexually-reproducing chironomid P. grimmii to be used as a standard ecotoxicological model insect. In a broader sense, this thesis aims to evaluate and mitigate some of the compromising effects of non-contaminant stressors (e.g. inbreeding and malnutrition) in ecotoxicology. These issues have often been overlooked or ignored, often to the detriment of a robust understanding of cause- effect relationships between contaminants and organismal stress effects. This thesis is divided into three experimental chapters and a general discussion chapter:

Chapter 2: A meta-analysis evaluating the relationship between aquatic contaminants and chironomid larval deformities in laboratory studies Aims: To compile, review and empirically analyse the literature of laboratory Chironomus deformity assays, to quantify the inconsistency of data and diagnose potential compromising/confounding factors. I performed a literature search of all laboratory studies using the databases Google Scholar, Scopus and Web of Science. This literature was reviewed to assess how many published experiments were potentially compromised by background stressors, and how many were potentially compromised by mortality in test larvae. I also performed meta-analyses assessing the inconsistency across assays of the most commonly tested

Page 20 contaminants: copper, zinc and lead, and also investigated all data for potential publication bias.

Chapter 3: A re-evaluation of chironomid deformities as an environmental stress response: sublethal toxicological assays and non-contaminant factors Aims: to clarify the causes of chironomid larval deformities, by eliminating the compromising factors that have limited previous work, and to consider novel, non- contaminant potential causal factors. I conducted sublethal assays, assessing the deformity-inducing effects of copper and the modern insecticide imidacloprid. I also investigated the non-contaminant stressor food limitation, and a non-stress associated factor: onset of metamorphosis (to determine whether deformities are in fact a natural phenomenon associated with the onset of pupation).

Chapter 4: The parthenogenetic cosmopolitan chironomid, Paratanytarsus grimmii, as a new standard test species for ecotoxicology: culturing methodology and sensitivity to aqueous pollutants. Aims: this chapter aimed to establish a basis for the use of P. grimmii as a model chironomid. Firstly, I aimed to develop a reliable, repeatable laboratory culturing method for this species. I then aimed to determine the sensitivity of P. grimmii to aqueous contaminants, to establish its baseline sensitivity (to copper, lead, cadmium, zinc and imidacloprid) and compare this with the sensitivity of model Chironomus species. This would enable the use of a non-inbreeding chironomid in studies, avoiding ID issues which are a widespread issue in assays using Chironomus.

Chapter 5: General discussion The findings of this thesis are summarised and discussed, along with potential future research directions.

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Page 30 Chapter 2 - A meta-analysis evaluating the relationship between aquatic contaminants and chironomid larval deformities in laboratory studies

Bryant S. Gagliardi, Vincent J. Pettigrove, Sara M. Long, & Ary A. Hoffmann (2016). A meta-analysis evaluating the relationship between aquatic contaminants and chironomid larval deformities in laboratory studies. Environmental Science & Technology, 50(23), 12903-12911.

Page 31 Article

pubs.acs.org/est

A Meta-Analysis Evaluating the Relationship between Aquatic Contaminants and Chironomid Larval Deformities in Laboratory Studies † † ‡ ‡ Bryant S. Gagliardi,*, Vincent J. Pettigrove, Sara M. Long, and Ary A. Hoffmann † Centre for Aquatic Pollution Identification and Management (CAPIM), BioSciences 4, School of BioSciences, The University of Melbourne, Melbourne, Victoria 3010 Australia ‡ Centre for Aquatic Pollution Identification and Management (CAPIM), Bio21 Institute and School of BioSciences, The University of Melbourne, 30 Flemington Road, Parkville, Victoria 3010 Australia

*S Supporting Information

ABSTRACT: Chironomid larval deformities have been widely investigated as an aquatic pollution toxicity end point. Field chironomid surveys often show a spatial association between contaminants and deformities, suggesting contaminants cause deformities. However, over 40 years of laboratory assays have not been able to confirm this causality. We therefore conducted a review of the literature and meta-analysis, in order to (A) assess whether trends across assays indicated dose−response effects, (B) characterize the consistency of results, and (C) investigate whether experimental issues and publication bias were contributing to inconsistency and/or reducing confidence in results. The experimental issues we investigated were extraneous nonchemical laboratory stressors (which may mask or interact with chemical effects), and mortality (which can confound deformity results). Our meta-analysis of the most commonly tested chemicals suggested dose−response effects for copper, but not lead or zinc. However, we also found substantial inconsistency across studies. Both mortality and extraneous stressors were potentially contributing to this inconsistency, reducing confidence in most published data. We observed no evidence of publication bias. We conclude that any causal link between contaminants and deformities remains uncertain, and suggest improved experimental and data reporting procedures to better assess this relationship.

■ INTRODUCTION the most recent chironomid deformity literature review was 3 Since the early 1970s, it has been observed that chironomid published over a decade ago, and there is no meta-analysis in the (Diptera: Chironomidae) larvae collected from polluted fresh- peer-reviewed literature, there is an opportunity to investigate water sites often show a higher frequency of deformities (of the trends across >40 years of published literature. mouthparts and/or antennae) than those collected from Meta-analysis can be used to evaluate cause-effect pollution 1,2 relatively unpolluted sites. These spatial associations between toxicity relationships across published studies.8 By analyzing the chemical contamination and deformities have led to the results of multiple studies, relationships can be evaluated across a frequently posed hypothesis that aquatic contaminants are causal broader range of concentrations than a single study, and indicate to chironomid deformities. Deformities have consequently been “ ” 7 widely investigated as a potential sublethal stress end point in the overall trends determined across studies. This approach − pollution toxicity studies.3 However, field results are subject to may be particularly useful in investigating dose response the influence of a variety of extraneous (i.e., nonchemical) relationships between chemicals and chironomid deformities, stressors,4 making it difficult to unequivocally attribute observed as the apparent variation between individual studies has deformities to contaminants. The evidence of chemical causality generated uncertainty as to overall dose−response effects.9 fi “ ” 3,5 produced in eld studies is hence largely circumstantial . Here we conduct a literature review and meta-analysis to explore  To further investigate causality, laboratory studies usually patterns in the literature on laboratory deformity studies. We first exposing Chironomus (the standard chironomid test genus) outline factors that might contribute to inconsistent and/or larvae to chemicalshave been conducted since the earliest fi observations of deformities.1 These studies have yielded unreliable results before considering the speci c aims of the apparently inconsistent results.5 This inconsistency may be study. indicative of experimental artifacts or experimental design issues, and reduces confidence in a causal relationship between Received: August 9, 2016 contaminants and deformities.6 Literature reviews6 and meta- Revised: October 18, 2016 analyses7 can serve to evaluate dose−response relationships, and Accepted: October 27, 2016 diagnose issues of concern across published bioassays. Given that Published: October 27, 2016

© XXXX American Chemical Society A DOI: 10.1021/acs.est.6b04020 Environ. Sci. Technol. XXXX, XXX, XXX −XXX Page 32 Environmental Science & Technology Article ■ SOURCES OF INCONSISTENCY AND probability of exhibiting deformities than the killed fraction UNRELIABILITY (which is typically eliminated from the analysis). A false negative may hypothetically arise if more sensitive larvaesusceptible to Results of different assays appear to be inconsistent, whether both chemically induced deformity and mortalityare killed by comparing tests of the same chemical (e.g., copper5), or the the chemical and hence excluded from analysis. This has been results of different chemical assays.5,9 Results have variously hypothesized to be the case in uranium16 and lead + zinc17 suggested positive, nonsignificant or even negative associations exposures, which yielded unexpected negative correlations between chemical concentrations and deformity frequency/ between chemical concentrations and deformity frequencies. severity.5 However, neither the “within-chemical” nor the “ ” fi A false positive may arise if a proportion of the chironomid across-chemical inconsistencies have been quanti ed. It is population is “naturally” deformed, that is, the apparently unknown whether results are systematically inconsistent, or if “deformed” morphology is simply a natural phenotype occurring only a small number of unrepresentative outlier results are giving at a certain frequency within the population. If individuals with a false impression of inconsistency. this morphology have a greater tolerance to the exposure While potential causes of inconsistent patterns have been chemical, they will better survive the exposure, hence the considered, they have rarely been empirically investigated. “ ” ff proportion of deformed larvae will increase with chemical Di erences between human observers in their designation of concentration. Such a correlation is actually a spurious larvae as either “deformed” or “non-deformed” has been 10 association between concentration and deformity frequency, as demonstrated as one potential source of inconsistency; and the chemical has not actually physiologically induced the the use of different Chironomus species (which can differ in their “ ” 11 deformity . This possibility (i.e., an association between chemical sensitivity ) may also contribute. Two additional deformities and mortality) may be suggested by results such as extraneous experimental factors have been suggested as those of Janssens de Bisthoven et al.,18 who observed a clear influential to chironomid deformity assays. These two factors, “ ” positive deformity-mortality correlation, in this case without background stressors and mortality of larvae, may contribute to observing such a clear chemical-deformity association. A false inconsistency if they are present at different magnitudes between ff fi positive may also arise even if there is no di erential chemical studies. They may also compromise (i.e., reduce con dence in) sensitivity between deformed and nondeformed larvae. For assay results, by potentially resulting in false positives or false instance, if deformed larvae have a lower survival overall because fi negatives. Another factor that can reduce con dence in published they are poorer competitors, an increase in mortality may lead to results is publication bias. These three issues (background a higher incidence of deformity simply because the level of stressors, mortality, and publication bias) are discussed hence- competition is reduced, irrespective of any inherent levels of forth. resistance. Confounding of deformity results by mortality can Background Stressors. Potential laboratory background/ therefore probably only be avoided by use of lower-than-lethal extraneous stressors (i.e., those other than the exposure exposure concentrations. chemical) include inadequate substrates, inbreeding, parental A high mortality rate in any treatment, as well as being the ff 3 e ects, and unknown stressors in substrates and food. result of chemically induced mortality, may be the result of  Experimental evidence supports at least two of these factors background mortality, which occurs independently of chemical 12 ff 13 inbreeding depression and parental e ects as potential concentration. As laboratory assays typically aim to eliminate all fl 4 in uences on deformity results. Given that the list of potential extraneous stressors, background mortality levels should 3 background stressors is long, the likelihood of them varying normally be low. However, when background mortality occurs, between studiesand hence contributing to result incon- mortality-susceptible larvae (which may have a higher or lower sistencyis high. Background stressors may also compromise probability of exhibiting deformity than those surviving) will results through additive, synergistic or antagonistic interactions usually be excluded from the analysis. In addition to selecting for with chemical toxicity, to increase/decrease the deformity mortality-resistant larvae, both chemically induced and back- 14,15 response, thereby risking false positive or false negative ground mortality reduce sample size. Both forms of mortality results. They may also lead to false negative results by “masking” hence also risk false negatives by reducing statistical power and chemical effects, that is, inducing a high deformity rate in control therefore the capacity of an experiment to detect a deformity 3 larvae, hence reducing the observed effect size between control effect (i.e., type II error). 15 and exposed larvae. Background stressors may also themselves Publication Bias. Another potentially compromising factor induce a deformity response.12 is publication bias. This occurs when the research that appears in Mortality. Mortality in assays can fall into two categories: the published literature is systematically unrepresentative of the mortality induced by the exposure chemical (“chemically- population of completed studies.19 Publication bias can come induced mortality”), and mortality induced by extraneous about when there is a bias toward publishing positive results that stressors (“background mortality”). The potentially distortive confirm or substantiate previous findings, or alternatively, when effects of mortality upon deformity results are complex, and are there is an increased likelihood of publishing negative results as outlined henceforth. an increasingly critical view of a field develops over time.20 It is a Although deformities are a sublethal end point, published potential problem within ecotoxicology.21 Publication bias in deformity assays often expose larvae to lethal chemical chironomid deformity studies has not been investigated concentrations.5 This has the potential to confound deformity previously. Although it may reduce confidence in published results, leading to false positive or false negative results. Larvae data,19 publication bias is not likely to contribute to result that survive a given concentration are likely to have numerous inconsistency. physiological differences from those killed by that concentration. This study presents a systematic, quantitative analysis of the “Likelihood of exhibiting deformities” may be one of these literature of laboratory chironomid deformity assays, in order to differences: the surviving fraction of the population (in which (A) assess whether trends across published assay results are deformities are typically quantified) may have a greater or lesser indicative of dose−response effects, (B) characterize the

B DOI: 10.1021/acs.est.6b04020 Environ. Sci. Technol. XXXX, XXX, XXX −XXX Page 33 Environmental Science & Technology Article inconsistency of results, and (C) investigate whether exper- for that concentration. For studies that reported deformity imental issues and publication bias are present. frequencies at several time points during an experiment (e.g., ref 27), we reported the total deformity frequency (i.e., that at the ■ MATERIALS AND METHODS termination of the experiment). Literature Search. We conducted an Internet-based search For each exposure treatment, we recorded dose magnitude as of the research literature (peer-reviewed journal articles and (exposure treatment concentration/control concentration) and academic text book chapters). In separate searches, we entered response magnitude as (exposure treatment deformity fre- the search terms “chironom* + deform*” and “chironom* + quency/control deformity frequency), following the meta- ’ 7 abnorm*” (to capture Chironomidae, chironomid, Chironomus, analysis methods of O Brien and Keough. The exact values of deformity, deformed, deform, deformation/s, deformities, deformity frequencies that were depicted graphically were not abnormalities and abnormal) into the search engine Google reported, so were estimated using the plot digitizing software Scholar and the databases Web of Science and Scopus, and PlotDigitizer (http://plotdigitizer.sourceforge.net/). Control imported all results using the reference management program chemical concentrations that were below the analytical limits 1 × 28 EndNote. The titles and abstracts of all English-language of detection (LOD) were recorded as /2 LOD, which also publications (published prior to 1 January 2016) were read, allowed a ratio to be taken. Analogously, deformity frequencies of and each publication investigating the incidence of chironomid 0 were replaced with half the maximum possible frequency in the larval deformities in laboratory chemical exposure experiments sample given the sample size: was saved. For each study, we recorded the test taxon used and = 1 freqdeformed larvae 1/2 + Treatments where the exposure chemical/s. ()nlarvae in treatment 1 Analysis 1: Meta-Analysis Assessing Potential Dose− LOD was not reported for chemical nondetects, or where Response Effects. Exposure chemicals analyzed in ≥3 nlarvae in treatment was not reported alongside zero deformity publications, using Chironomus as a test organism, were assessed frequencies, were therefore eliminated from Analysis 1. Each in Analysis 1. For each chemical, its concentration in each control data point in this meta-analysis represents the deformity and exposure treatment for each experiment were recorded. We frequency in a single exposure treatment within a single 7 also recorded whether experiments involved single or mixed- publication. chemical exposures. We recorded concentrations in each For each chemical, we then conducted Spearman’s correlation analyzed exposure matrix in ppm (i.e., mg/L for water analysis of deformity results to infer whether the overall data concentrations, mg/kg dry weight for sediment and tissue supported a dose−response relationship, which is a key indicator 29 concentrations). Although substrate contaminant concentrations of causality. For analysis and plotting, we took the ln of each are likely to be particularly relevant to the sediment-dwelling dose and response value.7 Therefore, positive ln(response) Chironomus,22 we noted that some studies did not report values corresponded to an increased deformity frequency in substrate concentrations, meaning we could not assess substrate response to chemical exposure, zero value to no response, and effects for these studies. For studies that reported both negative ln(response) values corresponded to a reduction in analytically determined and nominal concentrations, we deformity frequency in response to exposure. As there was recorded the analytically determined concentration as this is substantial spread in this data (in relation to the range of the more accurate exposure level. For studies that measured concentrations tested), ln-transforming the x-axis data meant it concentrations at several time points, we took the average could be plotted on a more meaningful scale. Potential dose− concentration determined (e.g., refs 23 and 24). We did not response effects for each chemical were inferred by a significant analyze results in which chemical concentrations in controls and/ (p < 0.05) positive correlation between ln(dose) and ln- or treatments were unreported. (response).7 Experiments using field-contaminated (as opposed to We also performed a sign test of this data to further test for laboratory-“spiked”) sediments were excluded from Analysis 1, potential causality. For each treatment in this data set, we with one exception, Di Veroli et al.23 This was because these recorded the number of positive, zero, and negative ln(response) studies (except for Di Veroli et al.23) compared deformity values. For each study (with 5 or more exposure treatments) we frequencies in relation to either “site from which sediment was tested whether the number of positive values was greater than the sampled” (i.e., control sediment versus polluted site sediment),25 number of nonpositive (i.e., zero plus negative) values. We also or “concentration of polluted site sediment” (i.e., a concentration performed this analysis at the “contaminant” level, pooling all gradient of polluted site sediment diluted with uncontaminated study values (n ≥ 5 treatments) for a given chemical. Potential sediment),26 rather than in relation to a specific contaminant chemical causal effects were inferred where the frequency of concentration. This made it difficult to analyze any chemical- positive responses was significantly higher (p < 0.05) than deformity dose−response relationship. nonpositive values. For each control and exposure treatment, we recorded the Analysis 2: Meta-Analysis Characterizing Inconsis- deformity frequency. Deformity “severity indices/scores” are tency of Results. To characterize inconsistency of results, we sometimes reported alongside deformity frequencies. These created a boxplot chart of ln(response) values from Analysis 1, indices are derived by applying a mathematical formula to separately creating boxplots for each of the most common deformity frequencies, such that treatments inducing “severe” exposure chemicals. In each case, results were considered deformities are scored as more deleterious than those inducing inconsistent if they, with the exclusion of potential outliers relatively “benign” deformities. As this distinction is subjective, [i.e., values outside the range [Quartile 1 − (1.5 × Interquartile and such indices have been shown to be redundant with range) to Quartile 3 + (1.5 × Interquartile range)], comprised deformity frequency results,9 such results were excluded from negative, zero and positive values, as implied by “inconsistency”.5 this analysis. For studies that investigated the deformity effects of Analysis 3: Literature Review Investigating Incidence a chemical concentration across several generations (e.g., refs 12 of Compromising/Inconsistency-Contributing Experi- and 18), we averaged the deformity frequency across generations mental Factors. We considered evidence of background

C DOI: 10.1021/acs.est.6b04020 Environ. Sci. Technol. XXXX, XXX, XXX −XXX Page 34 Environmental Science & Technology Article stressor, chemically induced mortality, and background mortality We therefore performed a funnel plot analysis of all treatments effects in the literature. The presence of these (or, the failure to for which deformity frequencies and sample size (i.e., number of satisfactorily eliminate their occurrence) was taken as evidence of larvae) was reported. We calculated LOR, that is, their potential contribution to inconsistency and reduced nndeformed larvae in exposure÷ non− deformed larvae in exposure fi ln , and its inverse stand- con dence in the results. This literature review analyzed results ()nndeformed larvae in control÷ non− deformed larvae in control at the “experiment” level, that is, each statistical analysis of an ard error,38 for each treatment. We plotted this data, and tested experimental result. Data for all exposure chemicals was analyzed for evidence of publication bias using Egger’s test for funnel plot in Analyses 3 and 4. asymmetry.39 Significant asymmetry was inferred if p < 0.05. As with Analyses 1 and 2, severity score/index results were not We also investigated whether “year of publication” of a study considered. Analyses of field-contaminated sediments were had a significant association with the magnitude of the deformity included, however, as Analysis 3 did not require the input of effect size (as LOR) in Chironomus studies. It has been suggested chemical concentrations. Also included, unlike Analyses 1 and 2, that such a “year effect” can be diagnostic of a tendency to favor were results that did not necessarily report deformity publication of results that either substantiate or refute previous frequencies, as long as statistical analysis was included. As per findings, which constitutes a publication bias.20 We conducted the previous analyses, we also excluded non-Chironomus Spearman’s correlation analysis of the data, and a potential results.30 publication bias was inferred if a significant (p < 0.05) positive To quantify the number of experiments which had apparently correlation (indicating a tendency to confirm previous findings) not eliminated background stressors (“Factor A”), we recorded or negative correlation (indicating a tendency to refute previous whether the deformity frequency in the control treatment was findings) was observed between year and LOR.20 All statistical reported and was <10%. We set 10% as the nominal maximum analyses in the present study were conducted in R.40 “acceptable” control deformity frequency as 1) it is the maximum acceptable control mortality rate for valid chironomid lethality ■ RESULTS assays according to the American Society for Testing and Literature Search. Our literature search found 41 studies 31 Materials International guidelines, hence is representative of a investigating chemical effects on chironomid deformity generally acceptable “control stress” level, 2) it is lower than 16%, frequency/severity in laboratory experiments (Supporting which is considered an unacceptably high control deformity Information, S1). All studies investigated effects in Chironomus 3 frequency by some researchers, and 3) it is generally concordant spp., with the exception of Dickman and Rygiel,30 which with field estimates of deformity rates of Chironomus populations investigated effects in Chironomidae. The latter study was inhabiting relatively unpolluted sites, though these can vary excluded from the following analyses in order to focus on 32 33 substantially, from 0% to as high as 20%. Chironomus. To quantify the number of experiments which had apparently Analysis 1. The most commonly assayed chemicals were not satisfactorily eliminated chemically induced mortality copper, lead, cadmium, and zinc (Table 1). Of these, the number (“Factor B”), we recorded whether the difference in mortality of studies reporting the parameters necessary for correlation rate between control and treatments was statistically analyzed analysis comprised four copper studies, seven lead studies, and and was not significant (p > 0.05). To further test for satisfactory three zinc studies. elimination of chemically induced mortality, and additionally, Potential dose−response effects were inferred for copper potential background mortality effects, for each analysis we (Figure 1A, ρ = 0.19, p = 0.03), but not for lead (Figure 1B, ρ = recorded whether the mortality frequency was reported for all 0.07, p = 0.31) or zinc (Figure 1 C, ρ = 0.09, p = 0.61). In the sign (control and exposure) treatments, and whether it was <20% in test analyses, the frequency of positive responses varied between each case (“Factor C”). Twenty percent mortality in controls is studies, however it was generally near or higher than 50% sometimes considered the maximum allowable for a valid (Supporting Information, Table S1). Significant study or chironomid lethality experiment (e.g. refs 34 and 35), so we contaminant-level effects were not found for any of the metals nominally set this as an “acceptably low” mortality rate. (p > 0.05), however only one zinc study had a sufficient number Analysis 4: Meta-Analysis Investigating Publication of treatments for analysis. Bias As a Compromising Factor. Publication bias can be Analysis 2. After the exclusion of outliers, ln(response) indicated by an association between sample sizes and outcomes results for copper, lead, and zinc spanned negative, zero and of experiments. In the absence of publication bias, outcomes of positive values, indicating a high level of inconsistency in experiments will be independent of sample size, as both large and experimental results (Figure 2). small experiments, resulting in either positive or negative Analysis 3. Two hundred and eighty-five statistical analyses, outcomes, will have an equal likelihood of publication. However, from 37 studies, were considered. For Factor A, the majority of an association between these two variables can indicate analyses appeared to avoid background stressors. However, these publication bias, often seen as a paucity of small experiments were still potentially common in published experiments: 70 out resulting in negative outcomes.36,37 This can be inferred by of 285 (i.e., 24.56%) analyses either did not report the deformity funnel plot analysis; for binomial data this can be investigated by frequency in controls, or reported a control deformity frequency plotting the log odds ratio (LOR, a measure of effect size) of a of >10%. The control deformity frequency was reported for 282 treatment against its inverse standard error (1/SE, an indicator of out of the 285 analyses; values ranged from 0 to 68.9% (average = sample size and hence precision; where a high 1/SE value 7.00%). corresponds to a large sample size).37 In the absence of Mortality was the more common factor potentially contribu- publication bias, this plot will form a symmetrical “funnel” ting to inconsistency. For Factor B, 263 out of 285 (i.e., 92.28%) shape around the mean LOR, with an increasing variability in of analyses either did not statistically compare mortality rate LOR as 1/SE decreases. Asymmetry about this mean may between control and treatments, or performed this comparison suggest publication bias, for example, a lack of data points with and observed a significant (p < 0.05) effect, indicating both low log odds-ratio and 1/SE values. nonavoidance of chemically induced mortality. Analysis for

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Table 1. Number of Studies Investigating Each Chemical for responses at lower doses. A negative response is where the Which Chemical Concentrations and Deformity Responses deformity frequency is reduced in response to chemical exposure, Are Reported meaning the interpretation of these patternsand inference of dose−response effectsis not straightforward. chemical number of studies Some ln(dose) values were <0 (Figure 1), meaning there was a Pb 10 higher chemical concentration in the control than in the Cu 5 experiment. Though unexpected, these values are still relevant Zn 4 to the dose−response analysis. For these values, corresponding Cd 3 negative ln(response) values were taken as “positive responses” 17α-ethynylestradiol 2 in the sign test analyses and Analysis 2 (see below). 4-nonylphenol 2 Meta-analysis data points can either represent a whole study, DDT 2 or a specific result within a study.42 In the latter situation, data Ni 2 points from the same study could be viewed as having a limited 2,4-D 1 independence. As such, the analysis may be viewed as having 210 Pb 1 relatively low statistical power (e.g., one may consider that n =3 acridone 1 in the zinc correlation analysis). The potential dose−response As 1 effects supported/not supported by Analysis 1 are therefore bisphenol A 1 subject to confirmation by further experimentation (see also the β -sitosterol 1 issues discussed below). chlorpyrifos 1 Analysis 2. To date, uncertainty about the causality of Cr 1 chemical contaminants to larval deformities has limited their dacthal 1 application in pollution monitoring and ecotoxicology; the DDE 1 consensus view is that laboratory results are too inconsistent to Di(2-ethylhexyl)phthalate 1 unequivocally demonstrate causal effects of chemicals upon − fenbendazole 1 deformities.9,18,24,41,43 45 Analysis 2 results are consistent with Hg 1 this perception: a substantial proportion of results suggest no imidacloprid 1 association, or even a negative association between chemicals and thiacloprid 1 deformities, even after the exclusion of potential outliers. tributyltin 1 However, there are methodological differences between U1studies that could weaken the detection of patterns in both xylene 1 Analyses 1 and 2. These include different researchers scoring ZnO microparticles 1 deformities, use of different Chironomus test species, inconsistent ZnO nanoparticles 1 categorization of deformity types, different chemical mixtures/ forms (e.g., metal salts) and concentration ranges, different Factor C found that a mortality rate of >20% in controls and/or exposure matrices, and differing experimental parameters (e.g., exposure treatments, or nonreporting of mortality rates, was also temperature); all of these may contribute to interexperiment common across analyses. This result was observed for 272 out of variability.5,10,11,23,44,46 Therefore, the inconsistency observed in 285 (i.e., 95.41%) of the analyses, potentially indicating Analysis 2 may be either due to these extraneous factors, or to the chemically induced and/or background mortality effects. A low natural variability in Chironomus deformity responses to percentage of analyses met all three criteria for studies without contaminants. If either of these situations is applicable, it potential compromising effects contributing to inconsistency: 4 suggests that a high frequency of deformities is not necessarily a out of 285 analyses (1.40%). All of the four “reliable” results reliable indicator of contaminant-induced stress, as deformity observed no significant association between chemical (β- frequencies do not consistently increase with concentration. In sitosterol, mercury or lead) concentration and deformity addition, the inconsistency of results may point toward “noise” in frequencies.41 the data. This may indicate that deformities are responding to Analysis 4. Four hundred and forty-six treatments (from 17 factors other than contaminant stress, that is, chemicals may not studies) reported sample size (i.e., the number of larvae be causal to deformities. Less “noisy” data may result from analyzed). The funnel plot of this data set appeared symmetrical deformity responses to other stimuli, which may suggest about the overall mean LOR, 0.46 (Supporting Information, nonchemical factors as causal to deformities. We have therefore Figure S1). This was confirmed by Egger’s test, which observed inferred that the relative frequency of zero and negative chemical- no significant departure from symmetry (t = −0.99, df =444, p = deformity associations may suggest experimental artifacts (see 0.32). This result indicates no publication bias based upon a Analysis 3) and/or a lack of causality.6 Negative associations can sample size analysis. Our temporal analysis also revealed no also suggest nonmonotonic relationships,2,47 and these may be publication bias, with no significant change in effect size detected supported by the frequency of negative ln(response) values at over time (ρ = 0.09, p = 0.052, Supporting Information, Figure low doses in Analysis 1 (Figure 1). However, this possibility S2). remains speculative, as ecotoxicological study designs typically involve too few exposure treatments and replicates for accurately ■ DISCUSSION diagnosing nonmonotonic/hormetic associations.48 Analysis 1. Our analysis of laboratory results observed Although our study assessed only laboratory results, field potential dose−response/causal effects for copper, but not lead results may be similarly inconsistent. For example, in an extensive or zinc. This provides some support for copper as causal to field study by Janssens de Bisthoven et al.,49 investigating 37 deformities. However, in the case of the correlation analysis, the deformity-chemical correlations, 24 were positive and 13 were pattern appears somewhat driven by a number of negative negative. We also noted that the LOR (i.e., effect size) data in

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Figure 1. Dose−response meta-analysis results for (A) copper, (B) lead, and (C) zinc. Spearman’s correlation analysis suggested dose−response effects for copper (ρ = 0.19, p = 0.03), but not lead (ρ = 0.07, p = 0.31) or zinc (ρ = 0.09, p = 0.61). Triangular data points represent substrate exposures, diamond-shaped points are for water exposures, circular points are for tissue concentrations.61,62,63

shows many values at or below x = 0, indicating a substantial number of zero and negative effects. This variability in this “general chemical” response may bring into question the suggestion that deformities are a general contamination indicator for field pollution assessments. Analysis 3. We found that background stressor and, more commonly, mortality effects, were potentially contributing to result inconsistency, reducing confidence in published data. By extension, there is a low confidence in the results of our Analyses 1 and 2. The potentially “compromising issues” we investigated have been suggested as potentially problematic either within chironomid deformity studies specifically,3,12,16,46 or in ecotox- icology more broadly.21 A low percentage (1.40%) of analyses met all three criteria for studies without potential background stressor or mortality Figure 2. Characterization of result inconsistency: boxplot of effects. All four of these “reliable” results observed no significant ln(response) data for Cu, Pb, and Zn exposures. Circular data points β “ association between chemical ( -sitosterol, mercury or lead) represent potential outlier values, gray line is the line of no chemical concentration and deformity frequencies.41 The null hypothesis effect”. of “no chemical effect” is therefore yet to be satisfactorily rejected in robust laboratory experimentation. Background stressors in Analysis 3testing responses across all chemicalsshowed most cases will be indicated by a high deformity frequency in substantial variability. Figure S2 (Supporting Information) controls.3,14 One background stressor, inbreeding depression,

F DOI: 10.1021/acs.est.6b04020 Environ. Sci. Technol. XXXX, XXX, XXX −XXX Page 37 Environmental Science & Technology Article can be eliminated by the use of apomictic parthenogenetic test have not reported mortality data, as potentially mortality- species;50 or minimized by use of large (>1000) laboratory affected. It is not our aim to criticize the authors of such studies, breeding population sizes and/or populations comprising field but to emphasize the importance of reporting (and statistically and/or mixed laboratory sources.14,51 Density/competition testing), mortality frequencies in sublethal ecotoxicological effects can be eliminated by exposing larvae individually in experiments. By comparison, sublethal chemical57,58 and UV separate test vessels.31 radiation59 levels in teratogenesis assays have observed dose- Although experimental evidence suggests mortality as a dependent deformity increases in amphibians. confounding experimental factor,16,18,46 it was common for The “acceptable” background mortality and control deformity studies to either not report mortality, fail to test for statistical rates we have applied in this study are nominal, and arguably differences in mortality between controls and treatments, or arbitrary. However, this argument can be made of any threshold publish results where a control-treatment mortality difference “acceptable” stress level. Nonetheless, these types of nominal was evident. For any experimental concentration inducing control values are routinely used as guidelines in ecotoxicol- significant mortality, it is difficult to determine whether the ogy,22,31,34,35 which can serve as coarse indicators of compro- observed deformity frequency is due to a direct chemical toxicity mised experimental conditions. More meaningful values might effect, or an indirect effect of mortality acting to increase/ be those known to have a biological relevance, for example a decrease deformity frequency within the surviving population.46 mortality rate known to result in a population-level decline, a The latter is an artifactual/indirect, rather than a causal, chemical deformity severity known to result in an individual-level association. Mortality effects could similarly be a compromising reduction in fitness, or a deformity frequency detected in a factor in field studies, but this is difficult to ascertain, given the well-replicated survey of unpolluted field sites. difficulty associated with quantifying mortality in wild field Additionally, any application of an “acceptable” control populations. While it is possible that “mortality-prone” and deformity frequency assumes that deformities are a “stress” “mortality-resistant” larvae have a similar likelihood of exhibiting (i.e., deleterious) response that is induced by environmental deformities, existing evidence suggests some difference.18 conditions, as opposed to other factors. This is yet to be clearly As deformities are a sublethal end point, investigations of demonstrated.9 Presently, little is known about the causes, chemical causality should involve sublethal concentrations,52 nature, and fitness consequences (at either the individual or over and above them being “environmentally relevant”. There is population level) of chironomid deformities. These are issues probably little value employing deformities to assess contami- requiring further research, potentially outside the context of nant-induced stress in field chironomid populations (e.g., in ecotoxicology. There is currently a bias toward chemical analyses, environmental monitoring programs) until chemical causality is with inbreeding depression12 and temperature25 to our knowl- established in controlled assays. Avoidance of mortality first edge being the only nonchemical factors considered in the requires that mortality be quantified in an assay. It is important laboratory. There may be value in investigating other stressors at that not only survival in all treatments should be comparable, but sublethal levels, for example, disease, starvation, and density that survival should be to the same larval instar (and statistically effects. tested for this, as in the β-sitosterol exposures in Vermeulen et Analysis 4. Analysis 4 revealed no evidence of publication al.41); as chironomid deformity results can also be potentially bias. While our temporal analysis suggested an upward trend in confounded by instar5 or developmental stage (e.g., if larvae are effects sizes over time, this effect was not statistically significant differentially lost to metamorphosis between treatments).26 (Figure S2, Supporting Information). The published data Second, avoidance of mortality requires that any mortality- therefore appear representative of all derived deformity data. inducing experiment be regarded as “questionable”, and repeated This also suggests that publication bias is not acting to reduce at lower (in many cases, lower than “environmentally relevant”) confidence in the published data. chemical exposure concentrations, and that this process be These inferences are limited to laboratory data, and field data repeated until a mortality difference in control-treatment have not been investigated for publication bias. We additionally comparisons is not detected, with a reasonable level of statistical note that sample size (number of larvae) was not reported for power to avoid type II error. In the first instance, lethality- over 200 treatments. This was either due to authors not reporting inducing chemical concentrations should not be used to the number of larvae chemically exposed/scored for deformities, investigate a causal relationship between that chemical and a or the number of larvae lost through either mortality, sublethal stress response. Perhaps only chemical concentrations metamorphosis, or failure to develop to the required instar was that have a significant mortality impact on a population are not quantified. Sample size is a key determinant of experimental capable of inducing deformities. However, in that case chemical precision. It is therefore important that, for each treatment, the effects would be difficult to differentiate from artifactual mortality exact number of deformity-scored larvae is quantified (such as in effects, except perhaps in cases where the control deformity Arambourou et al.60 and Di Veroli et al.23). frequency is zero. In summary, our meta-analyses revealed possible copper, but It is evident that authors have different views on the not lead or zinc dose−response effects upon chironomid importance of reporting mortality rates. In some cases, mortality deformities. However, we also revealed significant inconsisten- rates are explicitly quantified, in some they are referenced in text cies across published results. Potential background stressor without quantification, in some they are quantified and effects and mortality of larvae appeared to contribute to this statistically compared to control levels (e.g., refs 41 and inconsistency, and to a reduced confidence in most published 53−56), and in some studies they are neither mentioned nor data. Future experiments can better assess chemical-deformity quantified. The latter situation may sometimes indicate that no associations by eliminating these factors. Uncertainty regarding mortality has occurred, but in some such cases it is clear that these associations, along with the objective designation of mortality has occurred, such as studies that investigate effects at “deformity”,10 and the bearing of deformities on individual and LC50 levels. Because of this uncertainty, we have adopted a population fitness,9 currently limit the application of deformities conservative approach, where we considered any studies that as an ecotoxicological end point.

G DOI: 10.1021/acs.est.6b04020 Environ. Sci. Technol. XXXX, XXX, XXX −XXX Page 38 Environmental Science & Technology Article ■ ASSOCIATED CONTENT (11) Watts, M. M.; Pascoe, D. A comparative study of Chironomus riparius Meigen and Chironomus tentans Fabricius (Diptera:Chirono- *S Supporting Information midae) in aquatic toxicity tests. Arch. Environ. Contam. Toxicol. 2000, 39 The Supporting Information is available free of charge on the (3), 299−306. ACS Publications website at DOI: 10.1021/acs.est.6b04020. (12) Vogt, C.; Langer-Jaesrich, M.; Elsasser,̈ O.; Schmitt, C.; Van List of reviewed publications, sign test analyses, and results Dongen, S.; Köhler, H. R.; Oehlmann, J.; Nowak, C. Effects of of publication bias analyses. References: Supporting inbreeding on mouthpart deformities of Chironomus riparius under Information S1 List of reviewed publications Table S1 sublethal pesticide exposure. Environ. Toxicol. Chem. 2013, 32 (2), 423− Sign test analyses for experimental treatments Figure S1: 425. Funnel plot of experimental treatments for which samples (13) Servia, M. J.; Cobo, F.; Gonzalez, M. A. Incidence and causes of size (number of larvae) was reported. Figure S2:Scatter- deformities in recently hatched larvae of Chironomus riparius Meigen, 1804 (Diptera, Chironomidae). Fundam. Appl. Limnol. 2000, 149 (3), plot of year of publication of each deformity treatment 387−401. against log odds ratio (PDF) (14) Nowak, C.; Jost, D.; Vogt, C.; Oetken, M.; Schwenk, K.; Oehlmann, J. Consequences of inbreeding and reduced genetic variation ■ AUTHOR INFORMATION on tolerance to cadmium stress in the midge Chironomus riparius. Aquat. − Corresponding Author Toxicol. 2007, 85 (4), 278 284. * (15) Gagliardi, B. S.; Hoffmann, A. A.; Pettigrove, V. J. Inbreeding Phone +61383444331; e-mail: [email protected]. depression as a compromising factor in ecotoxicological assays. Integr. edu.au. Environ. Assess. Manage. 2016, 12 (3), 595−597. Author Contributions (16) Dias, V.; Vasseur, C.; Bonzom, J. M. Exposure of Chironomus The manuscript was written through contributions of all authors. riparius larvae to uranium: Effects on survival, development time, All authors have given approval to the final version of the growth, and mouthpart deformities. Chemosphere 2008, 71 (3), 574− manuscript. 581. (17) Martinez, A.; Moore, B. C.; Schaumloffel, J.; Dasgupta, N. Notes fi Induction of morphological deformities in Chironomus tentans exposed The authors declare no competing nancial interest. to zinc- and lead-spiked sediments. Environ. Toxicol. Chem. 2001, 20 (11), 2475−81. ■ ACKNOWLEDGMENTS (18) Janssens De Bisthoven, L.; Postma, J.; Vermeulen, A.; Goemans, We thank Dr Philippa Griffin for assistance with the layout of G.; Ollevier, F. Morphological deformities in Chironomus riparius Meigen larvae after exposure to cadmium over several generations. Figure 1, and four anonymous reviewers for helpful comments − − on the manuscript. This research was funded by the Centre for Water, Air, Soil Pollut. 2001, 129 (1 4), 167 179. (19) Rothstein, H. R.; Sutton, A. J.; Borenstein, M. Publication Bias in Aquatic Pollution Identification and Management, The Uni- Meta-Analysis: Prevention, Assessment and Adjustments; John Wiley & versity of Melbourne. Sons: West Sussex, 2006. (20) Jennions, M. D.; Møller, A. P. Relationships fade with time: A ■ REFERENCES meta-analysis of temporal trends in publication in ecology and (1) Hamilton, A. L.; Saether, O. A. The occurrence of characteristic evolution. Proc. R. Soc. London, Ser. B 2002, 269 (1486), 43−48. deformities in the chironomid larvae of several Canadian lakes. Can. (21) Wandall, B.; Hansson, S. O.; Ruden,́ C. Bias in toxicology. Arch. Entomol. 1971, 103 (03), 363−368. Toxicol. 2007, 81 (9), 605−617. (2) Warwick, W. F. 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I DOI: 10.1021/acs.est.6b04020 Environ. Sci. Technol. XXXX, XXX, XXX −XXX Page 40 Chapter 3 - A re-evaluation of chironomid deformities as an environmental stress response: Sublethal toxicological assays and noncontaminant factors

Bryant S. Gagliardi, Sara M. Long, Vincent J. Pettigrove, & Ary A. Hoffmann (in review). A re-evaluation of chironomid deformities as an environmental stress response: Sublethal toxicological assays and noncontaminant factors. Environmental Pollution

Page 41 Abstract

Larval deformities have been observed within both field and laboratory chironomid populations, and are thought to be connected to aquatic contaminant exposure.

However, chemical effects have not been directly related to deformities without potential confounding factors, particularly those due to mortality effects in laboratory assays. There is also a paucity of data on non-chemical causes of deformities. We therefore aimed to clarify factors associated with deformities, by generating sublethal toxicological data—assessing copper and imidacloprid effects—and considering non- chemical stressor (malnutrition) effects in the model chironomid Chironomus tepperi.

We also assessed whether deformities are associated with physiological changes induced during the onset of metamorphosis, in an attempt to better understand their origin. Our assays did not indicate an association between chemical and non- chemical stressors and deformities when tested at the sublethal level, and we found no association between deformities and the timing of metamorphosis. These results suggest a re-consideration of the utility of using deformities as an environmental stress response.

Keywords

Chironomidae; deformities; aquatic ecotoxicology; sublethal stress; pupation

Page 42 Introduction

Field-collected chironomid (Diptera: Chironomidae) larvae can exhibit deformities of the mouthparts and antennae (Janssens de Bisthoven and Gerhardt, 2005). Aquatic chemical contaminant levels and the incidence of chironomid deformities are often spatially associated in field samples, suggesting deformities indicate contaminant- induced stress (Janssens de Bisthoven et al., 1995). However, over 40 years of laboratory ecotoxicological experiments assessing this causal relationship (using the model genus Chironomus) have yielded inconsistent data and highlighted the potential impact of confounding effects (Vermeulen, 1995; Janssens de Bisthoven and Gerhardt, 2005; Salmelin et al., 2015; Gagliardi et al., 2016). This has resulted in uncertainty as to whether there is a causal relationship between chemicals and deformities, limiting the application of this endpoint in ecotoxicology.

One reason why laboratory work may not have clearly resolved the chemical- deformity relationship is that there are potentially confounding/compromising effects of mortality in toxicological assays (Gagliardi et al., 2016). Although deformity incidence represents a sublethal response, deformity assays often expose larvae to lethality-inducing chemical concentrations, and/or lethally stressful laboratory conditions. Dead larvae are eliminated from the analysis (due to their failure to reach the instar required for deformity scoring), so lethality-inducing experiments may select for “mortality-resistant” larvae rather than assessing deformity effects in the whole population (Polak et al., 2002). Mortality-resistant larvae may have a greater or lesser probability of exhibiting deformities than killed larvae, hence lethality- inducing experiments are potentially confounded by this factor (Martinez et al., 2004;

Gagliardi et al., 2016). Lethality-inducing assays risk producing false positive (if

Page 43 mortality-resistant larvae have a greater likelihood of exhibiting deformities) or false negative (if mortality-resistant larvae have a lower likelihood of exhibiting deformities) results. In other words, effects observed in lethality-inducing deformity assays may simply be the result of chemicals selecting for larvae with/without certain deformities, and do not necessarily indicate a mechanistic induction of deformities by contaminants. Lethality-inducing experiments are also confounded by density, as treatments in which larvae are killed will be under lower densities than those with high survival (Sanzo and Hecnar, 2006).

Potential mortality effects appear to compromise most published experiments on the relationship between deformities and contaminants, and may contribute to the fact that a substantial proportion of studies suggest no association between chemicals and deformities, or even a negative association (Martinez et al., 2004; Janssens de

Bisthoven and Gerhardt, 2005; Dias et al., 2008; Gagliardi et al., 2016). A recent meta-analysis found that > 95 % of published laboratory chemical assays either did not report mortality incidence across treatments, or reported deformity results for assays that induced mortality effects (Gagliardi et al., 2016). Agents causal to deformities should ideally be assessed in assays that eliminate mortality. Mortality effects can be excluded by ensuring survival (to the larval instar at which deformities are scored) is high across all control and exposure treatments, with no significant effects of treatments upon mortality frequency. This involves using sublethal exposure chemical concentrations for all treatments, ensuring laboratory conditions are favourable (as not to induce mortality), and reporting mortality levels across all experimental treatments.

Page 44 In addition to issues associated with mortality effects, another difficulty in clearly linking deformities to contaminant exposure is that deformities could be induced by other environmental stressors (in addition to, or to the exclusion of, contaminants).

Non-chemical effects on deformities have rarely been tested. There is a growing awareness of the importance of understanding the role of non-contaminant stressors in inducing ecotoxicological endpoint responses (Townsend et al., 2012; Dalzochio and Gehlen, 2016). Stressors such as malnutrition may induce chironomid deformities (de Haas et al., 2005; Janssens de Bisthoven and Gerhardt, 2005), however to our knowledge malnutrition-only deformity assays do not appear in the peer-reviewed literature.

It is also possible that abnormalities observed as larval “deformities” are due to normal developmental processes, in addition to/as opposed to being an environmental “stress” response. This possibility does not appear to have been previously considered, however larvae reared in uncontaminated laboratory conditions often exhibit deformities (Servia et al., 2000) which might reflect different patterns of development. Insects undergo many physiological and morphological changes during moulting and metamorphosis, and changes in the appearance of mouthparts and antennae may be associated with these processes. For example, as metamorphosis approaches, insects produce digestive enzymes which break down parts of the cuticle for “recycling” (Gilbert, 2000). Several chironomid deformities appear to be the result of cuticle dissolution, in particular the Kohn gap of the mouthparts [e.g. Figure 3D in Salmelin et al. (2015)], and the apparent “erosion” of terminal antennal segments [e.g. Figure 3F in Warwick (1985)]. Deformities in chironomid larvae may therefore also be the result of metamorphosis/moulting-

Page 45 associated biological processes, which occur as the larva transitions between larval instars, or towards the pupal stage.

This study aimed to clarify the causes of chironomid deformities, using the Australian model species Chironomus tepperi. Laboratory data on chironomid deformity incidence have previously been limited by confounding factors associated with mortality, and a paucity of non-toxicological data on deformities. To address this, we considered deformities under sub-lethal toxicant exposure, and considered non- chemical and developmental factors. We evaluated metal (copper) and insecticide

(imidacloprid) effects under sublethal conditions. These contaminants have previously been suggested as deformity-inducing, although results appear somewhat inconsistent (Vermeulen, 1995; Langer-Jaesrich et al., 2010), potentially due to mortality effects (Gagliardi et al., 2016). We also evaluated a non-contaminant stressor, food limitation, and a natural biological process, “onset of pupation” as deformity-inducing factors under sublethal conditions.

Materials and methods

Culturing of larvae for experiments

Chironomus tepperi were cultured as described in Jeppe et al. (2014) at 21 °C, using ethanol-rinsed toilet tissue as substrate, Tetramin as food, and “Modified Martins” medium (MM, hardness = 21–24 mg/L as CaCO3) as culture medium. For the experiment, second instar larvae were reared in 2000 ml beakers, as per Kellar et al.

(2014) and Mehler et al. (2017).

Page 46 Stress experiment

The experiment involved three types of stresses: two chemical exposure conditions

(imidacloprid and copper), a food limitation condition, as well as well-fed chemical- free controls. We used the metal salt CuSO4.5H2O (Univar, analytical reagent) and the granular commercial imidacloprid formulation Confidor (Bayer, 50g/kg active ingredient) in chemical exposures. In experiments, larvae were fed a combination of alfalfa, Spirulina, and α-cellulose powders. This combination has previously been shown to be nutritious for chironomid larvae, conferring high survival, emergence and reproduction in cultures (Gagliardi et al., 2015). A food suspension (FS) was freshly prepared for each feed, comprising 1 g of each of the three powders in 100 ml MM. Food suspension was aliquoted while being mixed with a magnetic stirrer, to ensure an even suspension of food particles.

Deformity results may be confounded by mortality [if mortality rates are significantly different between treatments (Gagliardi et al., 2016)], instar [if larvae in different treatments develop to different instars in significant numbers (Servia et al., 2002)], or life stage [e.g. if individuals in different treatments pupate/emerge in significantly different numbers (Hudson and Ciborowski, 1996; Gagliardi et al., 2016)]. To avoid these confounding factors, we aimed to ensure that survival of larvae to fourth instar did not differ across treatment concentrations within the experiment, and was ≥ 80% for all experimental treatments. We inferred this was the case where, at the termination of the experiment, ≥ 80 % of introduced larvae in each treatment were recovered (i.e. present and collected), all of which were alive and at the fourth instar stage. This endpoint, indicating collection/presence of surviving fourth instar larvae, is hereby abbreviated to “recovery”. We also ensured that, within each experimental

Page 47 condition, “recovery” rates were not significantly associated with stress (i.e. food/chemical level) in each condition (see Methods, Statistical analyses).

To determine sublethal concentrations, we conducted rangefinder experiments.

These involved three chemical concentrations (0.4, 4, and 40 mg/l for copper; 1.3, 13 and 130 µg/l for imidacloprid) and controls. The highest concentration that conferred recovery rates not significantly different to controls—and a recovery rate of ≥ 80%— was used as the highest exposure treatment in the experiment. These sublethal concentrations were 0.4 mg/l (copper) and 1.3 µg/l (imidacloprid). The subsequent experiment involved a series of five exposure treatments (dilution series = 0.6 ×) for each chemical. For food limitation treatments, we conducted a series of rangefinder experiments to determine an appropriate range of six FS volumes per feed, that would allow development of ≥ 80 % larvae from 2nd to 4th instar, with no significant association between recovery and food level. Chironomus larvae undergo three moults (hence four larval instars), and new deformities can arise in each instar

(Servia et al., 2002). It is therefore preferred that larvae in deformity experiments be exposed to contaminants for at least one moult. We therefore allowed larvae to undergo two moults during this experiment, to maximise exposure time to each of the environmental stressors and hence maximise the likelihood of inducing deformities.

Larvae were fed twice during rangefinders and experiments. We determined that the range 1600, 1040, 676, 440, 286, 186 µl FS per feed satisfied the recovery criteria.

We therefore used 1600 µl FS per feed in controls and chemical treatments, and

Page 48 1040 – 186 µl in food limitation treatments. Each stress condition therefore involved a gradient of sublethal five food/chemical levels plus controls.

Exposure vessels were 600 ml glass beakers (three replicate beakers for each treatment), containing 500 ml test solution (spiked MM for chemical exposures, unspiked MM for food limitation and control conditions), and 10 g acid-rinsed sand

(Chem-Supply, grain size = 300-350 µm) as substrate. Sand is commonly used as a substrate in chironomid studies assessing the effects of aqueous contaminants

(Mehler et al., 2011), and it has previously been shown that there are no significant differences in structural deformity frequencies between larvae reared on clay, paper, or sand substrates (Bird, 1997). Ten second instar larvae were introduced into each beaker with a plastic Pasteur pipette on Day 1 of the experiment. Beakers were incubated at 21 °C under a 16 h:8 h light:dark photoperiod and constant aeration.

Experiments ran for five days. On Day 3, physicochemical water quality parameters

[electrical conductivity (EC), dissolved oxygen (DO), pH; Mettler Toledo

SevenExcellence multi-probe meter] were measured in each beaker. Each beaker was also given a 50 % test solution change on Day 3, and fed on Days 1 and 3.

On Day 5, experiments were terminated by pouring beaker contents through a sieve

(pore size = 250 µm). Sieve contents were rinsed into a petri dish with MM and survival was recorded. Larvae were scored as dead if they did not move upon gentle probing with a pair of forceps. Larvae not recovered were presumed either dead/cannibalised, or emerged to adult stage. Larvae were then transferred into 1.5 ml Eppendorf tubes and preserved in 70-100% ethanol until deformity assessment.

Page 49 Chemical analyses

The chemical treatments involved preparation of two sets of exposure test solution: one for Day 1, and one for Day 3 (for the solution changes). Contaminant concentrations of all chemical test solutions were measured by commercial laboratories. Total copper in water was determined by Australian Laboratory

Services (Scoresby, Victoria), using inductively-coupled plasma mass spectrometry

(Perkin Elmer Elan 9000, US EPA Method 6020, NATA-accredited). Imidacloprid concentrations were determined by Advanced Analytical Australia (Kensington,

Victoria), using liquid chromatography–tandem mass spectrometry (compliant with

ISO/IEC 17025). These methods are described fully in Gagliardi et al. (2015). For statistical analysis of experimental results, chemical concentrations were recorded as the average concentration determined across the two solution preparations (Table

1).

Assessment of pupation-associated effects

To assess whether deformity incidence increased as larvae neared pupation, we conducted a 10-d developmental timecourse experiment. Thirty beakers were set up and incubated under the same conditions as the controls in the stress experiment.

On Days 3, 6, and 9, beakers were given a 50 % solution change and fed 1600 µl

FS. On Day 5, four replicate beakers were randomly selected, and larvae were analysed for recovery and deformities as per the stress experiment. These provided the control treatment for the timecourse experiment. Each day thereafter, 3-4 beakers were randomly selected, larvae were collected, and larvae analysed for recovery and deformity frequencies. Water quality (DO, pH, EC) in each beaker was

Page 50 determined at the time of termination. The experiment was terminated on Day 10, as recovery had reduced to < 80% (due primarily to pupation) at this time.

Deformity scoring

Larvae were slide-mounted under a light microscope, in Hoyer’s fixative, following

Gagliardi and Pettigrove (2013). Head capsules were observed under a compound microscope for deformities of the mentum, antennae and mandibles, adopting the scoring methods of Gagliardi and Pettigrove (2013). Deformities were “blind” scored to avoid biasing results (Salmelin et al., 2015). We tested for an abnormal number of mentum or mandible teeth, fusion or bifurcation of mentum or mandible teeth, Kohn gaps in the mandibles or mentum, abnormal numbers of antennal segments, and fusion of antennal segments. Previously, we had not considered antennae with terminal segments missing as deformed, as there was a possibility that these were broken during slide-mounting (Gagliardi and Pettigrove, 2013). However, we have since observed missing terminal antennal segments in live chironomid larvae under a light microscope, demonstrating this to be a true deformity. We therefore scored missing terminal antennal segments as deformities, as per other publications [e.g.

(Warwick, 1985)]. Although there is some disagreement between researchers as to the correct scoring of deformed menta, the particular mentum deformities we examined are relatively uncontroversial (Salmelin et al., 2015). We did not score as deformed menta or mandibles that appeared worn/broken due to mechanical damage/friction (Salmelin et al., 2015).

If a larva did not have—due to the arrangement of its features on the slide mount— all required features (mentum, both mandibles and both antennae) visible, it was

Page 51 scored as “deformed” if any of its visible features were deformed, but excluded from the analysis if none of its visible features were deformed. This was because it was impossible to determine the deformity status of the non-visible features. We also confirmed the instar of larvae by measuring mentum widths under the compound microscope with a calibrated graticule.

Statistical analyses

At the termination of each experiment we quantified recovery success (number of living 4th instar larvae present at the termination of the experiment, of the ten initially introduced larvae) for each beaker, and for effects among the treatments and controls within each condition (conditions = food limitation, copper, imidacloprid or pupation onset). We analysed for treatment (food/chemical level, or day) effects upon recovery with binary logistic generalized linear models (GLMs) (Szöcs and

Schäfer, 2015) to test whether deformity frequencies were not confounded by recovery (p > 0.05) (Gagliardi et al., 2016).

We calculated deformity frequencies (no. larvae with any deformed feature, per total no. recovered larvae) for recovered surviving fourth instar larvae in each beaker. We then tested for differences between conditions in the stress experiment (control, imidacloprid, food limitation, and copper, pooling food/chemical levels within each condition), by GLM analysis, analysing condition as a categorical factor. We then conducted GLM analysis within each condition, assessing for food/chemical treatment effects among the five treatment levels plus controls.

Page 52 For the timecourse experiment, we analysed for an effect of developmental day upon deformity frequency, also by GLM analysis. This analysis included all development days (Days 5 - 9) for which a recovery frequency of  80 % was observed. In all analyses, a significant association between any parameter and deformity frequency was inferred where p < 0.05. All statistical analyses were conducted in R (R Core

Development Team, 2016).

Results

Stress experiment

Measured chemical concentrations are reported in Table 1. Within each of the three conditions—copper, food limitation and imidacloprid—no treatment effects upon recovery frequency were detected (p > 0.05, Supporting information S1), indicating deformity results were not confounded by mortality, instar, or pupation (Gagliardi et al., 2016). In all treatments including controls, the number of surviving 4th instar larvae recovered was ≥ 80 % (i.e. ≥ 8 larvae per replicate mean), indicating laboratory conditions were favourable (Burton et al., 1996).

We observed deformities of the mentum, mandibles, or antennae in the stress experiment (see Figure 1 for examples) in some individuals, with an overall deformity incidence of 10 % (4 % in controls). The GLM analysis comparing each stress condition (grouped across treatment levels) indicated no significant deformity frequency differences between copper-exposed larvae and controls (z-score = 1.829, p = 0.07), nor between imidacloprid exposures (z-score = 1.146, p = 0.25) or food limitation conditions and controls (z-score = 0.385, p = 0.70) (Figure 2). In the comparison of food or chemical levels, no significant effects of copper (z-score = -

Page 53 1.558, p = 0.12), imidacloprid (z-score = -0.102, p = 0.92) or food limitation (z-score

= -1.382, p = 0.17) levels were observed, despite some variability in deformity frequency across treatments (Figure 3).

Developmental timecourse experiment

Deformities (Figure 1) were detected in this experiment, at a higher incidence than in the stress experiment (overall = 27%, control = 26%) (Figure 4). Recovery of larvae was ≥ 80% (i.e. ≥ 8 larvae per beaker) on Days 5 – 9 (Supporting information, S1).

We therefore analysed results across Days 5 – 9. Recovery success was not significantly associated with day of development across these five days (z-score = -

1.732, p = 0.08, Supporting information, S1), indicating deformity results were not confounded by instar, mortality or emergence. There was no significant effect of development day (z-score = 0.981, p = 0.33, Figure 4) upon deformities, suggesting that onset of pupation is not a driver of chironomid larval deformities.

Discussion

Stress experiment

Our results indicate that copper exposure, imidacloprid exposure, and malnutrition do not induce a higher incidence of chironomid larval deformities relative to controls.

Larvae in this experiment exhibited a low [i.e. < 10% (Gagliardi et al., 2016)] control deformity frequency (4%, Figure 3), suggesting an absence of extraneous deformity- inducing factors. These results are among the first chemical deformity assays published without apparent confounding by mortality, life stage, or extraneous stressors (Gagliardi et al., 2016). To our knowledge, the only other chemical assays meeting these criteria are the mercury, lead, and β-sitosterol assays of Vermeulen et

Page 54 al. (2000), which also observed no significant association between stressors and deformities. Together these results suggest that chironomid larval deformities may not be induced by environmental stress, and may not be suitable as an ecotoxicological endpoint.

There are trends in our results, however, that may warrant further examination.

Although not statistically significant, the results across the non-control treatments for the food limitation condition suggest an increase in deformity frequency as food increases. Additionally, the copper results may suggest a reduction in deformity frequency as copper concentration increases. These would be unexpected results, as they suggest that deformity frequency can decrease with increasing stress (i.e. lower food or higher copper). There may therefore be a complex association between stress levels and deformity frequencies, as also suggested by several previously published results, in which negative associations between stress level and deformity frequency have been observed in chironomids (Warwick, 1985; Martinez et al., 2004; Dias et al., 2008; Di Veroli et al., 2012).

While negative associations may simply be artefacts of high mortality rates (Gagliardi et al., 2016), they may alternatively point to the involvement of complex and indirect physiological and ecological factors, for example development rates. Chironomid development rate (usually measured as time to adult emergence) has a complex association with environmental stimuli. For example, synthetic pyrethroid pesticide exposure can slow development time under certain conditions (suggesting physiological stress) (Goedkoop et al., 2010), but accelerate it under others

(suggesting larvae may be seeking to “escape” contaminated environments by

Page 55 quickly metamorphosing and dispersing) (Boyle et al., 2016). A variety of non- chemical environmental factors also influence chironomid development rates (Ball and Baker, 1996; Olsen et al., 2003). Under favourable conditions, chironomid larvae may slow their development time to maximise feed time and build-up of energy resources, or may mature quickly in order to reproduce in a temporarily beneficial environment (Verberk et al., 2008). Perhaps larvae, stimulated to develop quickly by a variety of environmental factors (either stressful or favourable), invest less resources in forming larval head capsule features. They may instead invest more resources into fast development, and greater reproduction and/or dispersal capacity, resulting in the larval deformities observed in experiments. An association between development rate and deformities—as previously demonstrated in fish (Reimer et al.,

2017)—should be amenable to laboratory testing in chironomids. This would involve running test replicates in duplicate for each experimental treatment; one replicate would be terminated at the larval stage (for deformity scoring), and one allowed to develop to adulthood (to measure development rate, as emergence time), all under sublethal conditions.

Although our sublethal stress experiments and those of Vermeulen et al. (2000) do not support stressors as causal to deformities, there may be value in testing chemical combinations/types/concentrations not tested in these experiments, at sublethal levels. In addition, other potentially deformity-inducing stressors are yet to be tested at sublethal levels, such as larval density, temperature, and pathogens.

Parasites, for example, have been shown to induce morphological abnormalities in adult chironomids (Bhattacharya et al., 2014).

Page 56 Developmental timecourse experiment

In this experiment there was no effect of development day upon deformity frequency across the five pre-pupation days. This suggests that onset of pupation is not a physiological driver of chironomid larval deformities. Previous studies have suggested a positive association between time and deformity incidence in toxicological assays (Martinez et al., 2004), however these tests may not have been under sublethal/pre-pupation conditions.

There may also be patterns worth examining further in this experiment. Although not statistically significant, we observed an apparent increase in deformity frequency on the last analysed day (Day 9). We also observed that larvae approaching pupation underwent whole-body morphological changes, observable under a light microscope.

The most obvious of these was a thickening of the body section adjacent to the head capsule (the “thoracic” region). Presumably, any deformities associated with pre- pupation might manifest at this time, however we observed that larvae rapidly transitioned through this phase into pupae. As such, a 24-h sampling cycle was not sufficient to capture most individuals at this stage of development as, for many individuals, this entire stage had transpired between sampling events. To better test the hypothesis that this whole-body morphological change was associated with mouthpart and antennal deformities, a more time-intensive sampling protocol

(several times per day) might allow for a more comprehensive understanding of changes occurring during the pre-pupal phase. In addition, there may be interactive effects between environmental conditions, e.g. nutrition-stressed individuals may have a greater need to breakdown cuticle components for recycling during pre- pupation.

Page 57 It is also noteworthy that there is an apparent difference in control deformity frequencies between the developmental timecourse experiment (26%) and the stress experiment (4%). The relatively high frequency in this experiment suggests extraneous deformity-inducing factors influencing our results (Gagliardi et al., 2016).

It is not clear why there is such a difference between the two experiments, as experimental conditions were identical. It may be a result of the different physicochemical water quality parameters between the two (Supporting information,

S1), or may suggest that intrinsic biological factors, as yet unidentified, are the key driver of chironomid deformities. pH, DO and EC could therefore be tested as deformity-inducing factors [some field results have suggested EC and DO effects

(Odume et al., 2012), and EC associates with other stress responses in Chironomus tepperi (Hassell et al., 2006)]. Alternatively, deformity incidence may reflect sporadic changes in environmental or intrinsic biological factors—which are dynamic and unpredictable in their occurrence—and may therefore not make a suitable monitoring tool [which require a high repeatability of results (Burton et al., 1996; Harshman and

Hoffmann, 2000)].

We therefore suggest testing of water quality parameters such as EC, DO and pH— to our knowledge previously untested in the laboratory—as causal to deformities.

Previous work has also suggested a deformity effect of water hardness (potentially suggesting calcium or magnesium effects), so this could also be further tested

(Madden et al., 1995). We also suggest the importance of understanding the intrinsic biological factors involved in deformity formation. This may involve examining the genetic and biochemical differences between deformed and nondeformed larvae, within populations under unstressful laboratory conditions (often referred to as

Page 58 “background deformity” incidence). Inbreeding depression (Vogt et al., 2013), parental effects and “spontaneous dysgenesis” (Servia et al., 2000) are three intrinsic factors that have been suggested, there may be other intrinsic factors involved.

In summary, our experiments indicate no significant effect of imidacloprid, copper, food limitation, or onset of pupation upon deformity frequencies. These results suggest chironomid deformities may not be induced by stress, nor physiological changes associated with metamorphosis. These are among the first published results that have not introduced potentially confounding mortality effects in deformity experiments. In the case of the timecourse experiment, time-intensive sampling might be needed to further investigate developmental effects. We additionally suggest that future experiments consider the role of development rate, water quality parameters, non-chemical stressors, and biochemical and genetic factors in inducing deformities.

Acknowledgements

This study was funded by the Centre for Aquatic Pollution Identification and

Management, and the Holsworth Wildlife Research Endowment. We thank Dr

Philippa Griffin for assistance with the layout of Figure 3.

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Page 65 Table 1 Food amounts, n (replicate beakers), and nominal and measured chemical concentrations [mean (SD)] for deformity experiments conducted on Chironomus tepperi larvae

Experiment Condition Treatment Units n Measured concentration * Stress Non-stress Control - 3 Below LOD Copper exposure 0.05 mg/l 3 0.05 (0.005) 0.09 3 0.08 (0.01) 0.14 3 0.13 (0.01) 0.24 3 0.22 (0.01) 0.40 3 0.36 (0.02) Imidacloprid 0.17 µg/l 3 0.20 (0.02) exposure

0.28 3 0.30 (0.007) 0.47 3 0.49 (0.01)

0.78 3 0.79 (0.007) 1.30 3 1.35 (0.07) Food limitation 186 µl food suspension 3 - per feed 286 3 - 440 3 - 676 3 - 1040 3 -

Timecourse Onset of 5 (Control) Days 4 -

pupation 6 4 -

7 4 -

8 4 -

9 3 -

10 4 -

*Control stress experiment chemical concentrations were below limits of detection (LOD): < 0.001 mg/l copper, < 0.01 µg/l imidacloprid. Chemical, control and timecourse experiment beakers were fed 1600 µl food suspension per feed

Page 66

Figure 1 A) Non-deformed and B) deformed 4th instar Chironomus tepperi features. Top: mentum, centre: mandible, bottom: antenna

Page 67

Figure 2 Deformity proportions (jittered) across controls (n = 3) and different stress conditions (concentrations pooled, n = 15). Each datapoint represents the deformity proportion in a single replicate beaker

Page 68

Figure 3 Chironomus tepperi larvae deformity proportions (jittered) across stress treatments. Each datapoint represents the deformity proportion in a single replicate beaker, n = 3. Control chemical concentrations were below limits of detection (see Table 1), control and chemical treatment beakers were fed 1600 µl food suspension per feed

Page 69

Figure 4 Deformity proportions (jittered) in Chironomus tepperi over the five pre- pupation days of the timecourse experiment. Each datapoint represents the deformity proportion in a single replicate beaker

Page 70 Chapter 4 – The parthenogenetic cosmopolitan chironomid, Paratanytarsus grimmii, as a new standard test species for ecotoxicology: Culturing methodology and sensitivity to aqueous pollutants

Bryant S. Gagliardi, Sara M. Long, Vincent J. Pettigrove, & Ary A. Hoffmann (2015). The parthenogenetic cosmopolitan chironomid, Paratanytarsus grimmii, as a new standard test species for ecotoxicology: Culturing methodology and sensitivity to aqueous pollutants. Bulletin of Environmental Contamination and Toxicology, 95(3), 350-356.

Page 71 Bull Environ Contam Toxicol (2015) 95:350–356 DOI 10.1007/s00128-015-1578-5

The Parthenogenetic Cosmopolitan Chironomid, Paratanytarsus grimmii, as a New Standard Test Species for Ecotoxicology: Culturing Methodology and Sensitivity to Aqueous Pollutants

1 2 1 2 Bryant S. Gagliardi • Sara M. Long • Vincent J. Pettigrove • Ary A. Hoffmann

Received: 21 January 2015 / Accepted: 6 June 2015 / Published online: 12 June 2015 Ó Springer Science+Business Media New York 2015

Abstract Chironomids from the genus Chironomus are The standard international freshwater test insects for acute widely used in laboratory ecotoxicology, but are prone to (OECD 2011) and chronic (OECD 2010) ecotoxicology inbreeding depression, which can compromise test results. assays are species of the midge genus Chironomus (Dip- The standard Chironomus test species (C. riparius, C. tera: Chironomidae). According to protocols released by dilutus and C. yoshimatsui) are also not cosmopolitan, the Organisation for Economic Co-operation and Devel- making it difficult to compare results between geographic opment, the standard test species are C. riparius, C. dilutus regions. In contrast, the chironomid Paratanytarsus grim- (nee C. tentans) and C. yoshimatsui. However, laboratory mii is cosmopolitan, and not susceptible to inbreeding Chironomus spp. cultures often suffer from genetic depression because it reproduces asexually by apomictic impoverishment (Nowak et al. 2007b), reducing their parthenogenesis. However, there is no standardised cul- similarity with the wild populations they are supposed to turing methodology for P. grimmii, and a lack of acute represent (Brown et al. 2011). Genetic impoverishment toxicity data for common pollutants (metals and pesti- also leads to inbreeding depression, which can compromise cides). In this study, we developed a reliable culturing toxicological test results (Nowak et al. 2007a). In addition, methodology for P. grimmii. We also determined 24-h first Chironomus spp. are biologically different and regionally instar LC50s for the metals Cu, Pb, Zn, Cd and the endemic. Chironomus riparius is widely used in Europe insecticide imidacloprid. By developing this culturing and North America as a test species, C. dilutus is used in methodology and generating the first acute metal and North America and C. yoshimatsui is used in Japan. This imidacloprid LC50s for P. grimmii, we provide a basis for endemism means that test parameters (OECD 2011) and using P. grimmii in routine ecotoxicological testing. toxicity results (Watts and Pascoe 2000) are regionally specific according to the biology and toxicant sensitivity of Keywords Aquatic ecotoxicology Á Chironomidae Á the indigenous species used. The standard Chironomus Inbreeding depression Á Paratanytarsus grimmii Á species have a limited availability and ecological relevance Laboratory culture Á Apomictic parthenogenesis Á outside of their native ranges, necessitating ongoing work Chironomus on culturing methodologies and toxicant sensitivities for non-standard species. For example, C. striatipennis in Thailand (Somparn et al. 2010), and C. tepperi in Australia (Jeppe et al. 2014; Colombo et al. 2014) have recently been & Bryant S. Gagliardi [email protected] developed as regional model species. The chironomid Paratanytarsus grimmii may have 1 Centre for Aquatic Pollution Identification and Management advantages over Chironomus spp. for ecotoxicological (CAPIM), BioSciences 4, School of BioSciences, The testing. Genetic impoverishment in the laboratory is less University of Melbourne, Parkville, VIC 3010, Australia likely to be problematic in this species, as P. grimmii is 2 Centre for Aquatic Pollution Identification and Management clonal and has a naturally low genetic diversity (compared (CAPIM), Bio21 Institute and School of BioSciences, The University of Melbourne, 30 Flemington Road, Parkville, with other chironomids) across its native range (Carew VIC 3052, Australia et al. 2013). Any changes in genetic composition of 123

Page 72 Bull Environ Contam Toxicol (2015) 95:350–356 351 laboratory cultures can be easily tracked through moni- culture medium and commercially available foods. The toring clonal diversity. Furthermore, P. grimmii reproduces second aim was to determine 24-h LC50 values for first asexually by apomictic parthenogenesis (Porter 1971), instar P. grimmii to a number of chemicals, to document meaning inbreeding (and therefore inbreeding depression) the sensitivity of P. grimmii and thereby assess its suit- through mating of related individuals does not occur in this ability as a standard test species. species. Paratanytarsus grimmii is found in all continents except Antarctica (Langton et al. 1988; Hamerlı´k et al. 2011), so it could potentially be used in a global context. Methods and Materials Toxicity results obtained with P. grimmii, unlike endemic species, are likely to be relevant to aquatic ecosystems To initiate stocks for culturing trials, P. grimmii larvae and around the world. This species is also relatively easy to eggs were collected from Glynns Wetland, Warrandyte, source in the field, as it often dominates in terms of Victoria, Australia (-37.739869° latitude, 145.195388° abundance in sediment microcosm experiments within longitude) which is unpolluted (Pettigrove and Hoffmann Australasia (Pettigrove and Hoffmann 2005; Sharley et al. 2005). Plastic tubs, containing water and sediment from 2008) and Europe (Colombo et al. 2013). In addition, P. Glynns Wetland, were set up following Pettigrove and grimmii only consist of females (Porter 1971), meaning Hoffmann (2005). Chironomids were allowed to oviposit in toxicity results cannot be confounded by sex-specific dif- these tubs for 1 month, after which the tubs were returned ferences in pollution sensitivity as in Chironomus spp. to an outdoor shade house and covered with stockings. (Goedkoop et al. 2010). The development of standard Paratanytarsus grimmii are typically the dominant chi- methodologies for P. grimmii may therefore address many ronomid collected from these tubs. In subsequent weeks, P. concerns associated with use of Chironomus spp. in grimmii adults emerged and pupae became visible at the ecotoxicology. water’s surface, and were collected (the former with a While P. grimmii has been used previously in toxico- mechanical aspirator, the latter with a Pasteur pipette) and logical studies (e.g. Meier et al. 2000; Konishi et al. 2008), introduced into Perspex aquaria (described in Townsend there is no standard culturing methodology currently et al. 2012). available. Differing culturing methodologies can affect Initial attempts to culture P. grimmii involved the toxicity results when tests are carried out in different lab- method outlined in Jeppe et al. (2014) for culturing local C. oratories (Bradley et al. 1993). This is problematic, par- tepperi, however this resulted in few adults emerging. We ticularly when assays are used for regulatory purposes – then trialled a variety of culturing media, substrates and where reproducibility of results is essential (Soares and foods in all combinations (Table 1) using groups of 5 or Calow 1993). Published P. grimmii culturing methodolo- more adults to generate eggs for each trial. The conditions gies vary and are locally specific; for instance they cur- trialled were sourced from published descriptions of cul- rently require live algae as food and/or local natural water ture methods for a variety of freshwater organisms. For a as culture medium, both of which cannot be standardised trial, an aquarium was filled with 4 L of test culture (Bradley et al. 1993). In contrast, artificial culture media medium, and sufficient test substrate to cover the base of recipes and commercially available foods are prescribed the aquarium. The aquaria were kept at 21°C(±0.42 SD) for Chironomus spp. (OECD 2011). The development of a under constant gentle aeration and a photoperiod of standard protocol for P. grimmii should make it possible 16 h:8 h L:D. Adults were introduced to each aquarium for different laboratories to establish and maintain cultures using a mechanical aspirator and the colony was monitored and to establish similar standards when evaluating the for a generation (3–4 weeks) to quantify emergence. relative toxicity present in different environments. Once optimal culture conditions had been developed At present, limited acute toxicological data is available (see ‘‘Results and Discussion’’ section), they were used to for P. grimmii, particularly for the most commonly asses- produce adults whose offspring were used for toxicity tests. sed pollutants (metals and pesticides). Presently, only some The adult P. grimmii were transferred to a new aquarium chronic data for metals (Anderson et al. 1980) and a small with a mechanical aspirator. This aquarium contained the amount of pesticide data (Sato and Yasuno 1979; Hol- same medium as the culturing aquaria but was unaerated, combe et al. 1987; Meier et al. 2000) are available. These with a water column of *10 mm and no substrate. Adults data suggest that P. grimmii has a greater sensitivity to were allowed to oviposit overnight. Egg masses were pollutants when compared to other chironomids (Sato and subsequently collected with a Pasteur pipette and trans- Yasuno 1979; Eisler 1994) but more information is ferred to a petri dish containing culture media. This petri required, particularly for acute toxicity assays. dish was covered and incubated (21°C, 16 h:8 h L:D The first aim of this study was to develop a standard photoperiod) for 2 days, after which first instar (\24 h old) culturing methodology for P. grimmii, using artificial larvae were collected for exposures. 123

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Table 1 Culturing components Component Trialled condition References trialled in the development of a standard Paratanytarsus Freshwater culture medium ‘‘Modified Martins’’ Jeppe et al. (2014) grimmii culturing methodology ‘‘Sea salt’’ medium Giudice and Young (2010) Artificial soft water Hatakeyama and Yasuno (1981) M4 medium OECD (2011) M7 medium OECD (2011) Food ‘‘Tetramin’’ fish food Olsen et al. (2003) Alfalfa powder and spirulina powder Similar to Gersich et al. (1989) Substrate Granulated activated carbon Olsen et al. (2003) Shredded toilet tissue Jeppe et al. (2014) a-cellulose powder Marinkovic et al. (2011) ‘‘Artificial sediment’’ OECD (2010) Sphagnum peat moss OECD (2010) Sand OECD (2010) Kaolin Gaskell et al. (2007)

Following Be´chard et al. (2008), 24 h chemical expo- photoperiod were the same as in the culturing trials. The sures were carried out involving the common metal pol- average number of larvae per replicate was 9.55 lutants copper (as CuSO4Á5H2O, Ajax Finechem, (±1.06 SD), and there were 7–9 replicates per treatment. laboratory grade), cadmium (as CdCl2, Sigma Aldrich, All replicates within an experiment were exposed simul- technical grade), lead [as Pb(NO3)2, Ajax Finechem, ana- taneously rather than on separate days. lytical grade] and zinc (as ZnCl2, Ajax Finechem, labora- Test solutions were sampled prior to exposures and tory grade). We also included the common insecticide analysed by commercial laboratories. Metal concentrations imidacloprid [as the commercial formulation Confidor were determined by Australian Laboratory Services Global 200SC (Bayer), 200 g/L active ingredient], which has (Scoresby, Victoria) by inductively-coupled plasma mass previously been used in Chironomus spp. testing (e.g. spectrometry (Perkin Elmer Elan 9000, US EPA Method Stoughton et al. 2008). Modifications to the Be´chard et al. 6020, NATA-accredited). QA/QC procedures involved one (2008) procedure were as follows. Concentrations and method blank, one laboratory control sample, one matrix dilution series for each chemical (Table 2) were deter- spike and two duplicates for each group of 20 samples. mined in rangefinder experiments rather than using a log- Imidacloprid concentrations were determined by Advanced arithmic concentration range from 0 to 25 mg/L for all Analytical Australia (Kensington, Victoria) using liquid toxicants. For all experiments, the culture medium was chromatography–tandem mass spectrometry (compliant used as the exposure medium, and temperature and with ISO/IEC 17025). QA/QC procedures for these

Table 2 Nominal and analytical concentrations of chemicals used in toxicity tests with first instar P. grimmii, along with limits of quantitation (LOQ), dose–response models fitted to the data and 24-h LC50 (95 % CI) values Chemical Units LOQ Concentration Concentration Dose–response 24-h LC50 (95 % CI) type model C T1T2T3T4T5

Copper mg/L 0.001 Nominal 0 0.625 1.25 2.5 5 10 2 parameter log-logistic 1.33 (1.04–1.72) Analytical 0.004 0.538 1.08 2.06 4.03 8.18 Lead mg/L 0.001 Nominal 0 0.506 3.375 22.5 150 1000 3 parameter Weibull 27.85 (9.10–85.2) Analytical \LOQ 0.477 3.05 22.8 146 986 Zinc mg/L 0.005 Nominal 0 8.1 27 90 300 1000 2 parameter Weibull 96.73 (48.98–191.02) Analytical \LOQ 6.63 20.4 69.3 253 858 Cadmium mg/L 0.0001 Nominal 0 1.6 8 40 200 1000 2 parameter Weibull 70.64 (32.96–127.12) Analytical 0.0001 1.23 6 29.2 158 787 Imidacloprid lg/L 1 Nominal 0 8.1 27 90 300 1000 2 parameter Weibull 79.53 (48.84–129.51) Analytical \LOQ 6.9 25 83 270 890 C control, T (1–5) treatment

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Page 74 Bull Environ Contam Toxicol (2015) 95:350–356 353 analyses included a duplicate sample for every 10 samples particles), maintaining a substrate depth of *1 mm. Pho- analysed (acceptance criteria = 30 % relative percent dif- toperiod, aeration and temperature were the same as in ference), a matrix spike or laboratory control spike being culturing trials. performed one in every 20 samples with a recovery range Use of these culture conditions should increase the of 70–130 %, linearity with at least a 3 point standard utility of P. grimmii as an international standard test curve using a matrix matched standard, and a continuing organism for ecotoxicological studies. With our methods, calibration standard being run in the sequence for at least P. grimmii proved as simple to culture as C. tepperi. every 20 samples. An internal standard (Atrazine-d5) was However, the utility of P. grimmii also depends on the added to the samples and standards prior to analysis. sensitivity of this species to pollutants compared to Chi- Statistical analyses were conducted in R (R Develop- ronomus species. Our results present the first acute metal ment Core Team 2009) using the drc package (Ritz and and imidacloprid toxicity data for P. grimmii (Table 2) and Streibig 2005). Akaike’s Information Criteria were used to these can be compared to published data for Chironomus select the most parsimonious dose–response model species. Inbreeding depression has likely lowered some – among the 2-parameter (bounding the upper and lower LC50s in the literature for Chironomus spp. (Nowak et al. y-axis values at 1 and 0 respectively) and 3-parameter (as 2007a), but the P. grimmii results are likely to represent for the 2 parameter, plus an x-axis location parameter) log- reliable chironomid toxicity reference values. A compar- logistic, log-normal and Weibull models – for each ison of P. grimmii with Chironomus spp. acute LC50s chemical. The appropriate model was fitted to each dataset (Table 3) suggests that P. grimmii is more sensitive than C. (Table 2), and LC50s as well as 95 % confidence intervals riparius to copper, comparably sensitive as C. riparius to were then derived. zinc, and less sensitive than C. dilutus to both of these chemicals. Paratanytarsus grimmii appeared less sensitive than both Chironomus species to all of the other chemicals. Results and Discussion These results suggest that P. grimmii may be somewhat more pollution tolerant than C. riparius and C. dilutus.This The culturing conditions that generated emergence of at contrasts to a previous comparison between P. grimmii and least 10 adults involved Modified Martins (MM, culture C. yoshimatsui (Sato and Yasuno 1979) which suggested P. medium), a-cellulose powder (substrate) and a combina- grimmii as having sublethal pesticide EC50s several orders tion of spirulina powder and alfalfa powder (food). Three of magnitude lower than C. yoshimatsui. However, these aquaria were required to provide 10 or more adults at all species LC50s were all generated under different experi- times. The culturing methodology was standardised as mental conditions (exposure time, larval stage, water follows: aquaria contained a-cellulose powder as substrate hardness, feeding conditions), which can act to increase or (Sigma-Aldrich, product no. C6429, depth & 1 mm) and decrease LC50 values (Wang 1987; Williams et al. 1986;

4 L MM (hardness = 21–24 mg/L as CaCO3, pH 6.67, Nebeker et al. 1984). Species sensitivities should ideally be electrical conductivity = 180–220 lS/cm). Suspensions of compared under identical experimental conditions using alfalfa (‘‘Nutrigreen’’, http://www.nutrigreen.com.au) and outbred lines, but the standard test species do not occur in spirulina (‘‘Bioglan’’, http://www.bioglan.com.au) powder Australia. Moreover, while pollutant sensitivity is desirable were used as food, each consisting of 5 g of powder in in test species, it is generally not considered the most 100 mL MM. Cultures were fed 0.5 mL of alfalfa sus- important trait (Calow 1992), and test species with a gen- pension and 1 mL of spirulina suspension three times per eral sensitivity to all classes of pollutants have proven week. The a-cellulose provided habitat (used by larvae for elusive (Cairns 1986). Chironomids are standard test the construction of housing tubes), and may provide addi- insects because of their ecological relevance and tional nutrition (OECD 2011). To avoid fouling of culture amenability to manipulation under laboratory conditions medium by microbial activity, water quality was checked (Taenzler et al. 2007); these characteristics are also met by weekly, and the medium changed if one of the following P. grimmii. conditions applied: pH \6.5, dissolved O2 \80 % satura- The sensitivity of different P. grimmii clones to pollu- tion or [NH3] [1 mg/L. If electrical conductivity (EC) tants remains to be assessed. Previous research found no exceeded 220 lS/cm, reverse osmosis water was added apparent association between P. grimmii clonal type and until EC was returned to 180–220 lS/cm. Medium changes site pollution in the field (Carew et al. 2013). However, involved decanting the aquarium contents through nylon laboratory tests comparing clonal sensitivity still need to be mesh (pore size = 70 lm) and rinsing substrate and larvae carried out, along the lines of experiments with Daphnia back into the aquarium with 4 L fresh MM. After medium clones (Baird et al. 1989). Use of clones (whose pollutant changes, substrate was topped up with approximately sensitivities have been characterised) can allow for choice 10 mL fresh a-cellulose powder (to replenish lost fine between mixed clone experiments (where ecological 123

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Table 3 Comparison of experimental conditions (exposure time, instar, water hardness, feeding regime) used in toxicity tests and acute LC50s obtained in tests with Chironomus species and P. grimmii when exposed to different chemicals Chemical Units Species Exposure time (h) Instar Water hardness Fed? LC50 References (mg/L as CaCO3)

Copper mg/L P. grimmii 24 1 21–24 No 1.33 Present study C. dilutus 96 1 60–67 Yes 0.30 Nebeker et al. (1984) C. riparius 24 1 8 No 2.09 Be´chard et al. (2008) Lead mg/L P. grimmii 24 1 21–24 No 27.85 Present study C. riparius 24 1 8 No 0.61 Be´chard et al. (2008) C. dilutus 48 1 4.5–6 NR 2.68 Oladimeji and Offem (1989) Zinc mg/L P. grimmii 24 1 21–24 No 96.73 Present study C. riparius 24 1 8 No 25 Be´chard et al. (2008) C. dilutus 24 3 18–35 No 10.83 Khangarot and Ray (1989) Cadmium mg/L P. grimmii 24 1 21–24 No 70.64 Present study C. riparius 24 1 8 No 9.38 Be´chard et al. (2008) C. riparius 24 1 100–110 Yes 2.10 Williams et al. (1986) C. dilutus 24 3 18–35 No 23.25 Khangarot and Ray (1989) Imidacloprid lg/L P. grimmii 24 1 21–24 No 79.53 Present study C. riparius 24 1 NR NR 55.20 European Commission (2006) C. dilutus 96 3 140 Yes 5.40 Stoughton et al. (2008) NR not reported relevance is sought) or single clone experiments (where References precision is sought) (Forbes and Forbes 1993). The availability of a parthenogenetic test species will Anderson RL, Walbridge CT, Fiandt JT (1980) Survival and growth also allow new questions to be tackled, such as the role of of Tanytarsus dissimilis (Chironomidae) exposed to copper, cadmium, zinc, and lead. Arch Environ Contam Toxicol epigenetic effects in pollution adaptation (Vandegehuchte 9:329–335 and Janssen 2011). Clones are genetically identical, but Baird DJ, Barber I, Bradley M, Calow P, Soares AM (1989) The may differ in pollution responses if there are epigenetic Daphnia bioassay: a critique. Environmental bioassay tech- changes triggered by the environment. Such changes have niques and their application. Springer, Netherlands, pp 403–406 Be´chard K, Gillis P, Wood C (2008) Acute toxicity of waterborne Cd, been demonstrated in the cyclic parthenogen Daphnia Cu, Pb, Ni, and Zn to first-instar Chironomus riparius larvae. magna (Vandegehuchte et al. 2009), which reproduces Arch Environ Contam Toxicol 54:454–459 clonally under certain environmental conditions (Kleiven Bradley MC, Naylor C, Calow P, Baird DJ, Barber I, Soares A (1993) et al. 1992). Paratanytarsus grimmii, as an apomictic Reducing variability in Daphnia toxicity tests—a case for further standardization. Progress in standardization of aquatic toxicity parthenogen, reproduces clonally under all conditions, tests. Lewis, Boca Raton, pp 57–70 hence offspring will always be genetically identical to their Brown AR, Bickley LK, Le Page G, Hosken DJ, Paull GC, Hamilton mothers. For this reason, P. grimmii also presents as an PB, Owen SF, Robinson J, Sharpe AD, Tyler CR (2011) Are excellent test species for epigenetic studies in toxicological responses in laboratory (inbred) zebrafish repre- sentative of those in outbred (wild) populations?—A case study ecotoxicology. with an endocrine disrupting chemical. Environ Sci Technol In summary, we have developed a replicable and 45:4166–4172 simple culturing methodology for P. grimmii.Wehave Cairns J Jr (1986) The myth of the most sensitive species. BioScience also generated the first acute metal and imidacloprid 36:670–672 Calow P (1992) The three Rs of ecotoxicology. Funct Ecol 6:617–619 LC50s for this species. The widespread distribution of Carew M, Gagliardi B, Hoffmann AA (2013) Mitochondrial DNA this species and parthenogenetic mode of reproduction suggests a single maternal origin for the widespread triploid mean that P. grimmii may be useful for ecotoxicological parthenogenetic pest species, Paratanytarsus grimmii, but testing. microsatellite variation shows local endemism. Insect Sci 20:345–357 Colombo V, Mohr S, Berghahn R, Pettigrove VJ (2013) Structural Acknowledgments We thank Dr. Lisa Golding, Patrick Bonney, changes in a macrozoobenthos assemblage after imidacloprid Rebecca Reid and Katherine Jeppe for collecting the original P. pulses in aquatic field-based microcosms. Arch Environ Contam grimmii stocks. Funding was provided by the Centre for Aquatic Toxicol 65:683–692 Pollution Identification and Management. Colombo V, Pettigrove VJ, Golding LA, Hoffmann AA (2014) Transgenerational effects of parental nutritional status on Conflict of interest The authors declare no conflict of interest.

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Page 78 Chapter 5 – General Discussion

This thesis presents the first published empirical review of the chironomid deformity literature, and robust sublethal deformity toxicological data. Also presented is some of the first deformity data investigating non-chemical effects. Additionally, a replicable culturing methodology for the parthenogenetic chironomid P. grimmii is presented, which can facilitate international standardisation of aquatic insect ecotoxicological studies; and the avoidance of inbreeding effects confounding test results. The combined chapters of this thesis serve to highlight the importance of understanding and addressing non- chemical effects in ecotoxicological assays.

The meta-analysis revealed substantial inconsistency across published laboratory deformity data, and indicated much of this data to be potentially compromised/confounded by extraneous stressors and mortality of larvae. Almost all published deformity assays were potentially influenced by background stressor and/or mortality effects, introducing the risk of false positive and false negative results. These factors may be contributing to observed inconsistencies in deformity data where links are investigated with contaminant levels. Alternatively/additionally, the inconsistency of data may be suggestive of a lack of a causal effect of contaminants. Consistency of association between two variables is a Hill (1965) criterion for the inference of potential causality. For future deformity assessments, it is probably advisable for researchers to follow the Hill criteria for inferring potential causality, particularly when considering their own results in the context of all other published data. The Hill criteria are commonly advised for making robust inferences of causality in environmental pollution assessments (Vinken, 2013). Careful considerations to such criteria can minimise the risk of making spurious associations between potential causes and observed effect. In any case, the meta-analysis in Chapter 2 provides guidance for more robust assessment and reporting of stressor effects.

Page 79 The results of Chapter 2 also strongly suggest that a similar literature review is in order for field-based chironomid deformity assessments. Is there a tendency to favour publishing positive results in field studies? Is it common to observe “high” deformity frequencies in “control” sites, and “low” deformity frequencies in contaminated sites, as per the observations for laboratory studies (raising further doubts regarding chemical causality)? Furthermore, what actually constitutes “high” and “low” deformity frequencies probably bears much further consideration. For example, while I applied 10 % as a nominal “background” stress level in Chapter 2 [as per ASTM (2007)], a reading of the deformity literature suggests considerable differences in perceptions as to what constitutes high and low deformity levels. For example, a deformity frequency of 8.6 % was taken as evidence of significant stress in a contaminated site (Site ”Sb3”) in Beneberu and Mengistou (2014); yet a frequency of 10.83 % (“Station 1”) was considered sufficiently low to suggest unstressful contamination conditions in Arimoro et al. (2015). If 10 % (ASTM, 2007) or 20% (Burton et al., 1996) stress levels, as per convention, are indicative of an absence of contaminant-induced stress, it would be interesting to review the deformity literature (both laboratory and field) and quantify how many contaminant- exposed populations actually exceed the 10 (or 20) % threshold.

Taking on board these considerations, in combination with those relating to the significant “inter-researcher” variability in designating population deformity frequencies (Salmelin et al., 2015), it appears that there are still many fundamental concerns to be addressed in this area of study. These major hurdles will need to be overcome before chironomid deformities can be applied with any confidence in ecotoxicological evaluations. I suggest that tackling these issues should be prioritized of the continued publication of correlative field deformity investigations, which over the last 45 years have done little to empirically clarify the causal effects of contaminants.

Chapter 2 indicated that over 95 % of laboratory assays have potentially been confounded by mortality. Chapter 3 of this thesis avoids this issue and presents robust sublethal deformity data. In the stressor experiment of this chapter, I observed no significant deformity-inducing effects of copper, imidacloprid or

Page 80 food limitation, raising further doubts as to the role of stressors in inducing deformities. These results, alongside those of Vermeulen et al. (2000), are perhaps the only sublethal results published in the absence of evidence of background stressors (as indicated by a low control deformity rate). These data combine with those presented in Chapter 2 to further raise doubts as to the role of stressors in inducing deformities.

The timecourse experiment suggested that onset of pupation is not a driver of deformities, however, a more time-intensive sampling design may be required before fully ruling out this possibility. This was the first time to my knowledge that a non stress-associated factor has been considered as causal to chironomid deformities. There may be value in investigating other such factors, along with other stress-related factors/chemical concentrations (at sublethal levels) not assessed in this or the Vermeulen et al. (2000) study. Interestingly, there was an apparently different control deformity frequency between the stress experiment and timecourse experiment in Chapter 3. This was despite identical laboratory conditions between these experiments.

This result may suggest water quality effects. Water hardness, for example, is an important determinant of shell morphology in aquatic molluscs (Mackie and Flippance, 1983), and Chironomus larvae reared in culturing solutions of differing hardness by Madden et al. (1995) showed significantly different incidences of deformities. Additionally, water quality parameters such as salinity have been associated with a variety of stress responses in chironomids (Kefford et al., 2007). The differing control deformity frequency between studies may alternatively point to the involvement of variable, intrinsic physiological factors influencing deformity frequencies. These may be stress-associated, or they may be simply natural biochemical or genetic factors [such as different metabolic profiles (Silva et al., 2011) or different ploidy levels (Sadler et al., 2001)]. This is worth further investigation. Additionally, trends in the data may also be suggestive of a possible association between deformities and larval development rates, which should also be investigated (Reimer et al., 2017).

Page 81 Two questions remain when considering the utility of chironomid deformities in pollution monitoring: 1) are there causal effects of contaminants?, and 2) is this association consistent enough to generate repeatable environmental monitoring data? This thesis does not definitively answer Question 1 (though does strongly support a “no” answer), but the answer to Question 2 appears to be that “deformities are too unpredictable for this application” (e.g. Calow, 1992). At the least, the explicit quantification of the factors contributing to data inconsistency (use of different test species, environmental variables, intrinsic biological factors) is required before there can be progress in this area. Pursuit of these questions may require a change in the toxicant-focussed thinking that presently predominates chironomid deformity studies.

The establishment of deformities as an applicable endpoint will additionally need to consider their (empirically-established) merits against published guidelines for desirable biomarker traits (Lam, 2009). Whether they offer benefits over existing subcellular and individual-level endpoints is a key question. As for many individual-level endpoints, the important question whether chironomid deformities have any bearing upon individual or population fitness is also not yet known (Hämäläinen, 1999).

Even if deformities are eventually demonstrated as not useful for ecotoxicological applications, their cause/s may still to be of some academic interest. This thesis makes the case for consideration of a wider variety of factors (water quality, intrinsic biological processes etc.). Additionally, there is the philosophical question of what result/s (either from one study or a combination of studies) reasonably indicates “causality”, or more challengingly, a lack of “causality”? Does one positive result establish causality? Statistically, we might expect that one in 20 studies/experiments/concentrations gives a positive result by chance alone. In other words, a single positive result might be outweighed by a large number of negative results, in which case it does not provide sufficient evidence of “causality”. This is why I have referred to the Hill (1965) criteria, as this provides a robust set of guidelines for inferring potential “causality”. Consideration as to what constitutes a “likely-non causal”

Page 82 association with contaminants may be useful for evaluating of potential ecotoxicological endpoints in the future.

In the case of copper, lead and zinc, Chapter 2 showed that the number of positive deformity results was not greater than would be predicted by chance alone. It is therefore important for authors, when citing previous work, to consider whether it is a minority of studies that support causality, rather than selectively citing studies that support a causal effect of a given contaminant. Researchers need to be mindful of biases when appraising the overall literature (Nickerson, 1998). Furthermore, researchers also need to be mindful of biases when evaluating actual deformities (Salmelin et al., 2015). Hämäläinen (1999), for example, warned against the tendency for authors to grade deformities according to their apparent “severity”; applying “intuitive” (i.e. subjective) “horror scale[s]”. A lack of repeatability of results, raised previously in the biomedical (Ioannidis et al., 2009) and psychological (Gilbert et al., 2016) sciences, is becoming of increasing interest in the environmental and ecological sciences (Fidler et al., 2017).

The initial observations of chironomid deformities in the early 1970s were cause for reasonable scrutiny, as they suggested the widespread occurrence of teratogens in the environment, which can be of considerable concern for human and ecosystem health (Burkhart and Gardner, 1997). Pending further robust laboratory data, it may eventually be reasonable to temper these concerns, if data does not support larval deformities as indicative of the presence of teratogens.

The issues highlighted by Chapter 2, surrounding confounding of sublethal assays by mortality, may be of much wider concern than just deformity assessments. A brief appraisal of the ecotoxicological literature suggests that it is quite common for sublethal experiments to either not report mortality, or publish results where deformity effects are significant. With the apparent exception of investigations of contaminant-induced fluctuating asymmetry (Polak et al., 2002), the threat of confounding by mortality [also known as “survivorship bias” (Ho et al., 2012)] appears to have been largely overlooked in

Page 83 ecotoxicology. This thesis contends that, for assays investigating a causal link between contaminants and a sublethal response, inducing/not reporting mortality is unacceptable.

As it stands, the possibility of mortality effects simply selecting for individuals carrying/not carrying the sublethal trait of interest cannot be eliminated in any published lethality-inducing assay. Drawing mechanistic (rather than simply correlative) links between contaminants and sublethal responses in any such assay therefore remains problematic. Determining the response level in “mortality-prone” individuals (i.e. those killed in the course of the experiment) at the conclusion of an experiment does not resolve this problem. This is because these individuals—through dying—will have “shutdown” their physiological processes during the experiment, so are not comparable to individuals whose metabolism was active for the entire duration of the exposure.

The confounding effects of mortality may explain some anomalous published ecotoxicological results, for example, the observation of sublethal E(ffect)C50/I(nhibitory)C50 concentrations higher than LC50 concentrations for exposed laboratory populations (Maul et al., 2008). Such data may be a potential result of the seemingly common practice of conducting a mortality (e.g. LC50) assay, and then determining sublethal response/s in the residual surviving organisms, risking a confounding by mortality effect. In such an experiment, the sublethal effect is not being measured in “the population”, but rather, “the surviving proportion of the population, after a mortality event”.

Given that sublethal means “lower than lethal”, it would be unusual for a truly sublethal toxic concentration to be higher than a lethal one. As a brief example of the potentially widespread nature of mortality issues, the following ecotoxicological guideline methods do not make explicit recommendations regarding avoiding lethality in sublethal assays across all exposure concentrations: Davies and Freeman (1995), OECD (2010), and OECD (2012). A literature review of this area is therefore required for appraising the extent of the problem. This thesis suggests that, to avoid survivorship bias, experiments with sublethal endpoints that induce a mortality effect need to be re-conducted

Page 84 at lower, truly sublethal concentrations. As it stands, it appears that survivorship bias may be a major problem within ecotoxicology.

Paratanytarsus grimmii can be cultured using internationally replicable methods based on commercially available components, with guidelines presented in Chapter 4. Methods described for C. tepperi (Jeppe et al., 2014) and other Chironomus species did not appear to be appropriate for P. grimmii.

The toxicological results in Chapter 4 suggest P. grimmii may have a higher chemical tolerance than the model Chironomus species. Chemical sensitivity is a desirable trait in model species, although it may be difficult to have a generally “sensitive model species” (Cairns, 1986). In any case, my sensitivity comparisons were based upon literature values for Chironomus species LC50s. Direct assays with multiple species may provide a more robust means of comparing species, as differences in experimental conditions between assays can have substantial effects. Such comparative assays were not possible for me to achieve, as the standard Chironomus test species—C. riparius, C. dilutus and C. yoshimatsui—do not occur in Australia (highlighting the importance of using a cosmopolitan species). Furthermore, formal statistical comparisons with literature Chironomus LC50s were not achievable, as such comparisons require reporting of either 95 % confidence intervals or standard errors alongside LC50s (Wheeler et al., 2006). The errors for literature Chironomus LC50s reported in Chapter 4 were variously reported as standard deviation, standard error, confidence intervals, or in some cases, no error value was reported.

A starting point for Australian laboratories might involve assays comparing C. tepperi and P. grimmii. Whether inbred or outbred C. tepperi lines are used is an important question, and depends somewhat on the phrasing of the experimental hypothesis: i.e. whether the aim is to compare P. grimmii sensitivity with the “true” sensitivity of C. tepperi, or with that of cultured C. tepperi populations. The complex issue of whether the sensitivity of inbred cultured Chironomus spp. populations is representative of the “true” respective species’ sensitivity is further discussed in Appendix 1.

Page 85 In any case, the ability of P. grimmii to avoid inbreeding effects may outweigh concerns regarding its apparent contaminant tolerance. Inbreeding depression (ID) has been suggested as a compromising variable in Chironomus deformity assays (Vogt et al., 2013) (although it is not clear whether ID was tested at sublethal levels in this study, meaning these assays may also have been compromised by survivorship bias). Observation of deformities in P. grimmii, in any case (a chironomid model in which ID is impossible), will suggest that ID may not be a/the driver of deformities in Chironomidae. Paratanytarsus grimmii may therefore make an excellent model for examining larval deformities. This species does exhibit several types of deformities, including those typically observed in Chironomus (personal observation, from pilot studies).

As well as determining the sensitivity of P. grimmii in comparison to Chironomus, the question of whether different [apparently regionally endemic (Carew et al., 2013)] clonal lines differ significantly in sensitivity is also an important avenue for future research. This is analogous to previous research conducted on Daphnia magna clonal lines. Differential sensitivity of D. magna lines to contaminants has been established (Baird et al., 1991). This in turn has opened up other areas of research, such as the investigation of genetic versus non-genetic modes of organismal contaminant tolerance (Barata and Baird, 1998). Field surveys suggest that P. grimmii clonal lines may be comparable in their sensitivity (Carew et al., 2013), but it is important to investigate these questions in the laboratory.

Another important step in the development of P. grimmii as a model organism is the development of standard assay methods, similar to those for Chironomus species (US EPA, 2000; OECD, 2010). Like Chironomus, P. grimmii larvae are sediment-dwelling/burrowing/ingesting (personal observation), so in this regard will almost certainly make an appropriate contaminated sediment toxicity testing species, as well as being suitable for aqueous contaminant testing. A challenge when working with P. grimmii is the relatively small size of larvae: they are much smaller than larvae of Chironomus species. The experiments described in Chapter 4 involved transferring of first instar larvae with a micropipette, under a light microscope. This would probably be too laborious for many applications,

Page 86 particularly those requiring high throughput, such as environmental sediment monitoring programs [e.g. Kellar et al. (2014)]. Similarly to Chironomus, early (probably first and second instar) P. grimmii are pale/white in colour. This, in combination with their small size, makes them difficult to see with the naked eye (especially against non-contrasting backgrounds), limiting their application in any test that cannot be conducted without the use of a light microscope (e.g. beaker-based sediment tests, from which it would be near-impossible to recover burrowed early instar P. grimmii). Therefore, for these types of applications, fourth instar P. grimmii larvae will probably have to be used, depending on whether these are sufficiently sensitive to contaminants. It is therefore important to determine the contaminant sensitivity of each larval instar, as testing in Chironomus species has shown contaminant sensitivity to decrease with increasing instar (Williams et al., 1986). This will require standardised conditions (time, larval densities, food amounts) that can reliably produce larvae of a desired instar for experiments. It is challenging to determine the larval instar by observing body size, so head capsule size/mentum width should be measured to determine this (McCauley, 1974).

In summary, routine use of P. grimmii appears be achievable with further work. This work should investigate comparative sensitivity assays with C. riparius, C. dilutus, C. yoshimatsui and other regional model species. Also required is determination of the contaminant sensitivity of different P. grimmii clonal lines and instars, and development of standard contaminant testing methodologies.

In conclusion, this thesis aimed to evaluate and tackle several of the non- contaminant factors that appear to have been compromising results in aquatic insect ecotoxicology. A thorough analysis of the chironomid deformity literature is presented, clarifying published trends and suggesting improved analytical and reporting methods for more robust deformity work. Robust, sublethal chironomid deformity data is also presented, which raises doubts as to the links between stressors and deformities, and does not suggest metamorphosis-associated effects. Finally, the ability to generate non ID-compromised, internationally applicable aquatic insect data—by use of P. grimmii in assays—is also presented. This thesis highlights the importance of understanding and

Page 87 differentiating the responses of organisms to contaminant and non-contaminant stressors in ecotoxicological assays, and dedicating effort to avoiding these extraneous factors where possible.

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Page 92 Appendix 1 – Discussion Paper: Inbreeding depression as a compromising factor in ecotoxicological assays

Bryant S. Gagliardi, Ary A. Hoffmann, & Vincent J. Pettigrove (2016). Inbreeding depression as a compromising factor in ecotoxicological assays. Integrated Environmental Assessment and Management, 12(3), 595- 597.

Page 93 Integr Environ Assess Manag 12, 2016—PM Chapman, Editor 595 accuracy required to ensure that aquatic life guidelines for Se depression (ID), which occurs when rare deleterious recessive are applied appropriately. alleles become expressed when they occur as homozygotes as a consequence of inbreeding; and through the loss of heterozy- REFERENCES gote advantage. Inbreeding depression can result in a loss of fi Beatty JM, Russo GA. 2014. Ambient water quality guidelines for selenium organism tness. technical report update. Victoria (BC): BC Ministry of Environment. 254 pp + Inbreeding depression therefore acts similarly to pollutants, appendices. thermal extremes, diseases, and habitat degradation in that it [CCREM] Canadian Council of Resource and Environment Ministers. 1987. induces deleterious effects. It is potentially compromising to Canadian water quality guidelines. Ottawa (ON): Task Force on Water Quality the aims of laboratory ecotoxicology because it is a stressor that Guidelines. is enhanced, rather than eliminated, in a laboratory setting. 50 FR 46902. 1985. National primary drinking water regulations: Volatile synthetic Although ID has been studied extensively in conservation – organic chemicals. Federal Register 50:46902 46906. biology, it has received relatively less attention in ecotoxico- Janz D, Brooks ML, Chapman PM, DeForest D, Gilron G, Hoff D, Hopkins B, logy (Brown et al. 2009). Research suggests, however, that ID McIntyre D, Mebane C, Palace V, et al. 2010. Selenium toxicity to aquatic is of concern. For some stress endpoints, ID interacts with organisms. In: Chapman PM, Adams WJ, Brooks ML, Delos CG, Luoma SN, Maher WA, Ohlendorf HM, Presser TS, Shaw DP,editors. Ecological assessment chemical toxicity, such that inbred organisms can have a of selenium in the aquatic environment. Pensacola (FL): SETAC Press. p 141–231. greater chemical sensitivity (and hence endpoint response) Kentucky Energy and Environment Cabinet. 2013. Update to Kentucky water than outbred animals (Nowak, Jost, et al. 2007; Brown et al. quality standards for protection of aquatic life: acute selenium criterion and 2009). For other endpoints, ID has been shown to induce tissue-based selenium chronic criteria. Frankfort (KY): Kentucky Energy and certain endpoint responses independently of a chemical Environment Cabinet. 35 pp + appendices. response (Nowak, Jost, et al. 2007). [USEPA] US Environmental Protection Agency. 2003. Analytical feasibility support Some ramifications of ID for laboratory ecotoxicology have document for the six-year review of national primary drinking water been previously discussed by Brown et al. (2009). This regulations. Washington (DC): USEPA. EPA-815-R-03-003. discussion focused on the interactive effects of ID and chemical [USEPA] US Environmental Protection Agency. 2009. Analytical feasibility support document for the second six-year review of existing national primary drinking toxicity and on their potential to compromise ecological risk water regulations. Washington (DC): USEPA. EPA 815-B-09-003. assessments (ERAs). The authors point out that use of ID- [USEPA] US Environmental Protection Agency. 2015. Draft aquatic life ambient sensitized organisms in ERAs—which aim to extrapolate water quality criterion for Selenium—Freshwater 2015. Washington (DC): laboratory findings to protect natural ecosystems—may lead to USEPA. EPA 822-P-15-001. overly protective/conservative ecological toxicity estimations. They suggest strategies for minimizing inbreeding in cultures, such as the use of large (>1000) reproducing population sizes or paired breeding programs. They also point out the importance of quantifying the relationship between genetic INBREEDING DEPRESSION AS A COMPROMISING diversity levels and endpoint stress levels, and of identifying the FACTOR IN ECOTOXICOLOGICAL ASSAYS “critical” inbreeding levels above which fitness is compromised Bryant S Gagliardi,*y Ary A Hoffmann,y and Vincent J Pettigrovey (possibly the critical inbreeding coefficient of F ¼ 0.33) (Brown yCentre for Aquatic Pollution Identification and Management, et al. 2009). School of Biosciences and Bio21 Institute, University of In addition to sensitizing organisms (Nowak, Jost, et al. Melbourne, Parkville, Victoria, Australia 2007; Brown et al. 2009), ID may also “mask” chemical effects. *[email protected] DOI: 10.1002/ieam.1766 This masking could occur when both ID and chemical exposure independently (without interaction) induce a particular endpoint response. Masking is illustrated in a Laboratory ecotoxicological assays investigate causal relation- hypothetical scenario as follows. The increase in stress response ships between toxicants and biological stress responses. In field apparently associated with Chemical X in Figure 1A is experiments, it can be difficult to differentiate “true” pollutant relatively small (the stress level at 40 mg/L being only 1.1 effects from those induced by extraneous stressors (e.g., thermal that of the control treatment), suggesting a low toxicity. extremes, diseases, habitat degradation) (Chapman 1995). Figure 1B, however, differentiates ID-induced endpoint These extraneous stressors can covary with pollution, are responses from chemically induced endpoint responses for sometimes undetectable, and may have complex interactions this experiment and reveals the observed effects to be mostly with pollutants involving additivity, synergy, or antagonism. due to ID. Another experiment, exposing outbred organisms Adding to this difficulty is the fact that most endpoints are to the same Chemical X concentrations (Figure 1C), induces general stress—rather than pollution-specific—indicators. the same chemical effects. However, the absence of inbreeding Laboratory assays, by eliminating extraneous stressors, provide effects in this experiment means that the observed stress an important tool for investigating cause–effect relationships. response is much greater (10 that of the control treatment at Organisms for assays are usually reared in “in-house” 40 mg/L), revealing a much greater toxicity. Masking in ERAs cultures. These cultures represent captive colonies. Captive may result in an underestimation of, or failure to detect, colonies are often subject to a reduction in genetic diversity chemical toxicity, producing insufficiently protective toxicity (relative to wild populations). This has been shown to be the values that risk ecosystem health. case for ecotoxicological cultures; genetic impoverishment is Both sensitization (for endpoints subject to an ID- common in cultures of the test aquatic insect Chironomus chemical interaction) and masking (for ID-affected end- riparius (Nowak, Vogt, et al. 2007) and fish Danio rerio (Coe points not subject to an interaction) may be problematic for et al. 2009). Inbreeding, defined as reproduction between ERAs. Some standard assays require a sufficiently high fitness closely related individuals, is a common consequence of in control organisms for a test to be considered valid (such as genetic impoverishment. It often results in inbreeding a minimum control survival frequency in a mortality assay).

Page 94 596 Integr Environ Assess Manag 12, 2016—PM Chapman, Editor

Figure 1. Hypothetical experimental results for inbreeding depression–compromised (A, B) and outbred test organisms (C) exposed to a toxic chemical. Toxicity results are “masked” by inbreeding depression in A and B.

This can help avoid situations where “background” stressors Furthermore, the influence of ID on most endpoints remains compromise the sensitivity of test organisms. However, unquantified. this requirement will not screen out ID-compromised Given that ID is potentially widespread in cultures (Nowak, animals for endpoints for which there is an ID-chemical Vogt, et al. 2007; Coe et al. 2009), we suggest that interaction, but no effect of ID alone. In these cases, there is consideration be given to the possibility that many published likely to be no detectable effect of ID in control treatments, toxicity values may be influenced (either inflated or reduced/ but a substantial ID effect at high chemical exposure masked) by ID effects. We also advocate that researchers, concentrations. when planning and publishing experiments, clearly define the While ERAs aim to protect ecosystems, ecotoxico- specific aims of their study (e.g., an ERA aimed at protecting logical assays more broadly aim to determine “true” chemical ecosystems, an academic study aimed at investigating “true” toxicity effects. Whether chemical effects observed in chemical effects), and that they consider and discuss the ID-compromised organisms are “true” toxicity effects, or extent to which ID may compromise these aims. Research is rather those of “combined stressor” effects (such as those also required to determine which specific endpoints, at what that may be observed in chemical exposures of disease- level of inbreeding, are compromised by ID (Brown et al. compromised organisms), is a matter for further consideration. 2009). We therefore suggest that the issue of ID in ecotoxicology is Ideally, inbreeding in cultures is to be minimized as much one that requires further consideration and research. Despite as possible, by use of large laboratory population sizes, previous discussion (Brown et al. 2009), ID continues to be paired breeding programs, and/or regular “refreshment” of given insufficient attention. Assays often source organisms cultures with individuals from the field or several external from cultures whose population size is unreported, or are likely laboratories (Nowak, Jost, et al. 2007; Nowak, Vogt, et al. inbred (due to their single source laboratory population, small 2007; Brown et al. 2009). Reporting of culture sources, laboratory population size, or many generations of captivity). approximate population size, and refreshment frequency

Page 95 Integr Environ Assess Manag 12, 2016—PM Chapman, Editor 597 will allow the reader to make broad judgments about the olfaction can affect both the individual fish and the entire fish likelihood of ID influencing results. Reporting of inbreeding population. levels (as inbreeding coefficient F) and/or genetic diversity The paired olfactory organ consists of 2 olfactory cavities at levels of cultures will allow more accurate judgments to be the dorsal part of the snout. Each is connected with the made. environment via 2 openings, allowing the surrounding water to Additionally, ID in cultures can be completely avoided by flow through. The multilamellar olfactory epithelium is use of obligate parthenogenetic test species that reproduce by located on the ground of the cavity. It contains the olfactory apomixis, such as the aquatic test insect Paratanytarsus grimmii receptor neurons (ORNs), which are responsible for odor (Gagliardi et al. 2015). Asexual reproduction involves no detection. These neurons are only separated by a thin layer of mating, so inbreeding is impossible. Though nonobligate test mucus from the surrounding water. In this extremely exposed parthenogens such as Daphnia magna can avoid inbreeding situation, they can interact easily with odor molecules such as during their asexual phases, they are subject to inbreeding prostaglandins or amino acids present in concentrations down during their sexual phases. Apomictic parthenogens have no to 1012 M and 109 M, respectively (Hara 1992). Because the sexual phase and a mitotic reproductive cycle. Heterozygosity surrounding water not only contains natural odorants but also a is hence maintained across generations. Apomicts, through variety of contaminants, it is likely that these are able to clonal reproduction and maintenance of heterozygosity, interact with the ORNs and thereby might have an adverse completely avoid ID and hence are an attractive option in effect on fish olfaction. ecotoxicology. Some metals and pesticides are known to impair the Acknowledgement—The authors thank W Tyler Mehler for olfactory system of fish. For example, following 30 min helpful comments on a draft of this manuscript. exposure to 10 mg/L Cu, the olfactory response in Coho salmon (Oncorhynchus kisutch) was reduced by 70% relative to REFERENCES the control (Baldwin et al. 2003). Despite the high importance and sensitivity of olfaction for fish, to date there are Brown AR, Hosken DJ, Balloux F, Bickley LK, Lepage G, Owen SF, Hetheridge MJ, comparatively few studies on this topic; the underlying Tyler CR. 2009. Genetic variation, inbreeding and chemical exposure— Combined effects in wildlife and critical considerations for ecotoxicology. Phil mechanisms of olfactory toxicity are still mainly unknown. Trans R Soc B 364:3377–3390. Most of these studies were conducted in Canada and the Chapman J. 1995. The role of ecotoxicity testing in assessing water quality. Aust J United States, and native salmonid species were used as model Ecol 20:20–27. organisms. They are relatively big fish, require a lot of space, Coe T, Hamilton P, Griffiths A, Hodgson D, Wahab M, Tyler C. 2009. Genetic their development is slow, and their breeding is seasonal and variation in strains of zebrafish (Danio rerio) and the implications for complicated. ecotoxicology studies. Ecotoxicology 18:144–150. In Europe, olfactory toxicity in fish is a scarcely regarded Gagliardi BS, Long SM, Pettigrove VJ, Hoffmann AA. 2015. The parthenogenetic issue. The zebrafish (Danio rerio), a small freshwater cyprinid, cosmopolitan chironomid, Paratanytarsus grimmii, as a new standard test is a common model organism used in medical and biological species for ecotoxicology: Culturing methodology and sensitivity to aqueous research all over the world. The use of zebrafish embryos is pollutants. Bull Environ Contam Toxicol 95:350–356. Nowak C, Jost D, Vogt C, Oetken M, Schwenk K, Oehlmann J. 2007. receiving increasing attention, because they are considered a Consequences of inbreeding and reduced genetic variation on tolerance to replacement method for animal experiments. Up to 120 h cadmium stress in the midge Chironomus riparius. Aquat Toxicol 85: postfertilization (hpf) they are classified as nonprotected 278–284. and, thus, nonregulated stages in the sense of EUDirective Nowak C, Vogt C, Diogo JB, Schwenk K. 2007. Genetic impoverishment in 2010/63/EU (EU 2010). Danio rerio is inexpensive, easy to laboratory cultures of the test organism Chironomus riparius. Environ Toxicol handle, and produces a large amount of translucent eggs Chem 26:1018–1022. throughout the year. Its genome has been completely sequenced and a variety of molecular and genetic methods have been established. The olfactory system of zebrafish is typical for many teleost fish species; further, their olfactory fi OLFACTORY TOXICITY IN FISH—WHY WE SHOULD signal transduction is similar to mammals. Zebra sh larvae are CARE ABOUT IT able to detect odors and even distinguish between different odor classes as early as 72 hpf (Vitebsky et al. 2005). y y y Sina Volz,* Sabrina Schiwy, and Henner Hollert The aim of our research strategy is to contribute to the yDepartment of Ecosystem Analysis, RWTH Aachen University, establishment of zebrafish as a model organism for olfactory Aachen, Germany *[email protected] toxicity, using it to study the mechanisms underlying DOI: 10.1002/ieam.1777 impairment of olfaction. In addition, we want to identify and evaluate suitable test endpoints for olfactory toxicity in fish that could be applied in the risk assessment of The olfactory system is of great importance for fish as it chemicals. mediates a variety of essential activities and behaviors. Apart The impact of selected substances on endpoints at different from its role in foraging, it also plays an important part in levels of organization will be investigated. Whenever possible, homing, for instance salmon follow olfactory cues to find experiments will be conducted on both adult zebrafish and their home water (Hasler et al. 1978). Olfactory alarm cues their early life stages. A comparison of the collected data will released from injured fish trigger a predator avoidance reveal whether experiments with (the nonprotected) zebrafish behavior in fish of the same species. In male fish, the larvae are suitable alternatives for animal investigations with perception of female pheromones leads to an increase in adult fish. plasma testosterone followed by an enhanced production of So far, the pesticides chlorpyrifos and permethrin have been expressible milt (Moore and Lower 2001). An impairment of used as test substances. For chlorpyrifos, there are already

Page 96 Appendix 2 – Supplementary Material to Chapter 2

Page 97 Supporting information

A meta-analysis evaluating the relationship between aquatic contaminants and chironomid larval deformities in laboratory studies

Bryant S. Gagliardi1,*, Vincent J. Pettigrove1, Sara M. Long2, Ary A. Hoffmann2

1 Centre for Aquatic Pollution Identification and Management (CAPIM), BioSciences 4,

School of BioSciences, The University of Melbourne, Victoria 3010 Australia

2 Centre for Aquatic Pollution Identification and Management (CAPIM), Bio21 Institute and

School of BioSciences, The University of Melbourne, 30 Flemington Road, Parkville, Victoria

3010 Australia

*Corresponding author: [email protected]

Includes list of reviewed publications, sign test analyses, and results of publication bias analyses (10 pages, one table, and two figures)

S1

Page 98 S1 List of reviewed publications

Arambourou, H.; Beisel, J.N.; Branchu, P.; Debat, V., Exposure to sediments from polluted rivers has limited phenotypic effects on larvae and adults of Chironomus riparius. Sci. Total

Environ. 2014, 484, 92101.

Arambourou, H.; Beisel, J. N.; Branchu, P.; Debat, V., Patterns of fluctuating asymmetry and shape variation in Chironomus riparius (Diptera, Chironomidae) exposed to nonylphenol or lead. PLoS ONE 2012, 7 (11).

Arambourou, H.; Gismondi, E.; Branchu, P.; Beisel, J. N., Biochemical and morphological responses in Chironomus riparius (Diptera, Chironomidae) larvae exposed to leadspiked sediment. Environ. Toxicol. Chem. 2013, 32 (11), 25582564.

Bird, G. A.; Schwartz, W. J.; Joseph, D. L., The effect of 210Pb and stable lead on the induction of menta deformities in Chironomus tentans larvae and on their growth and survival. Environ. Toxicol. Chem. 1995, 14 (12), 21252130. den Besten, P. J.; Naber, A.; Grootelaar, E. M. M.; van de Guchte, C., In situ bioassays with

Chironomus riparius: Laboratoryfield comparisons of sediment toxicity and effects during wintering. Aquat. Ecosyst. Health Manage. 2003, 6 (2), 217228.

Dias, V.; Vasseur, C.; Bonzom, J. M., Exposure of Chironomus riparius larvae to uranium:

Effects on survival, development time, growth, and mouthpart deformities. Chemosphere

2008, 71 (3), 574581.

Dickman, M.; Rygiel, G., Chironomid larval deformity frequencies, mortality, and diversity in heavymetal contaminated sediments of a Canadian riverine wetland. Environ. Int. 1996,

22 (6), 693703.

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Page 99 Di Veroli, A.; Goretti, E.; Paumen, M. L.; Kraak, M. H. S.; Admiraal, W., Induction of mouthpart deformities in chironomid larvae exposed to contaminated sediments. Environ.

Pollut. 2012, 166, 212217.

Gremyatchikh, V. A.; Tomilina, I. I.; Grebenyuk, L. P., The effect of mercury chloride on morphofunctional parameters in Chironomus riparius Meigen (Diptera, Chironomidae) larvae. Inland Water Biol. 2009, 2 (1), 8995.

Hamilton, A. L.; Saether, O. A., The occurrence of characteristic deformities in the chironomid larvae of several Canadian lakes. Canad. Entomol. 1971, 103 (03), 363368.

Hudson, L. A.; Ciborowski, J. J. H., Teratogenic and genotoxic responses of larval

Chironomus salinarius group (Diptera: Chironomidae) to contaminated sediment. Environ.

Toxicol. Chem. 1996, 15 (8), 13751381.

Janssens de Bisthoven, L.; Huysmans, C.; Vannevel, R.; Goemans, G.; Ollevier, F., Field and experimental morphological response of Chironomus larvae (Diptera, Nematocera) to xylene and toluene. Neth. J. Zool. 1997, 47 (2), 227239.

Janssens de Bisthoven, L.; Postma, J.; Vermeulen, A.; Goemans, G.; Ollevier, F.,

Morphological deformities in Chironomus riparius Meigen larvae after exposure to cadmium over several generations. Water, Air, Soil Pollut. 2001, 129 (14), 167179.

Janssens de Bisthoven, L.; Vermeulen, A.; Ollevier, F., Experimental induction of morphological deformities in Chironomus riparius larvae by chronic exposure to copper and lead. Arch. Environ. Contam. Toxicol. 1998, 35 (2), 248256.

Kwak, I. S.; Lee, W., Mouthpart deformity and developmental retardation exposure of

Chironomus plumosus (Diptera : Chironomidae) to tebufenozide. Bull. Environ. Contam.

Toxicol. 2005, 75 (5), 859865.

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Page 100 Kosalwat, P.; Knight, A. W., Chronic toxicity of copper to a partial lifecycle of the midge,

Chironomus decorus. Arch. Environ. Contam. Toxicol. 1987, 16 (3), 283290.

LangerJaesrich, M.; Koehler, H.R.; Gerhardt, A., Assessing Toxicity of the Insecticide

Thiacloprid on Chironomus riparius (Insecta: Diptera) Using Multiple End Points. Arch.

Environ. Contam. Toxicol. 2010, 58 (4), 963972.

LangerJaesrich, M.; Koehler, H.R.; Gerhardt, A., Can mouth part deformities of

Chironomus riparius serve as indicators for water and sediment pollution? A laboratory approach. J. Soils Sediments 2010, 10 (3), 414422.

Lewis, J. W.; Morley, N. J.; Ahmad, M.; Challis, G. L.; Wright, R.; Bicker, R.; Moffitt, D.,

Structural changes in freshwater fish and chironomids exposed to bacterial exotoxins.

Ecotoxicol. Environ. Saf. 2012, 80, 3744.

Madden, C.; Austin, A.; Suter, P., Pollution monitoring using chironomid larvae: What is a deformity? In Chironomids: from genes to ecosystems, Cranston, P. S., Ed. CSIRO: East

Melbourne, 1995.

Madden, C. P.; Suter, P. J.; Nicholson, B. C.; Austin, A. D., Deformities in chironomid larvae as indicators of pollution (pesticide) stress. Neth. J. Aquat. Ecol. 1992, 26 (24), 551557.

Martinez, E. A.; Moore, B. C.; Schaumloffel, J.; Dasgupta, N., Effects of exposure to a combination of zinc and leadspiked sediments on mouthpart development and growth in

Chironomus tentans. Environ. Toxicol. Chem. 2004, 23 (3), 662667.

Martinez, A.; Moore, B. C.; Schaumloffel, J.; Dasgupta, N., Induction of morphological deformities in Chironomus tentans exposed to zinc and leadspiked sediments. Environ.

Toxicol. Chem. 2001, 20 (11), 247581.

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Page 101 Martinez, E. A.; Moore, B. C.; Schaumloffel, J.; Dasgupta, N., Morphological abnormalities in Chironomus tentans exposed to cadmium—and copperspiked sediments. Ecotoxicol.

Environ. Saf. 2003, 55 (2), 204212.

Martinez, E.; Wold, L.; Moore, B.; Schaumloffel, J.; Dasgupta, N., Morphologic and growth responses in Chironomus tentans to arsenic exposure. Arch. Environ. Contam. Toxicol. 2006,

51 (4), 529536.

Meregalli, G.; Ollevier, F., Exposure of Chironomus riparius larvae to 17αethynylestradiol:

Effects on survival and mouthpart deformities. Sci. Total Environ. 2001, 269 (13), 157161;

Meregalli, G.; Pluymers, L.; Ollevier, F., Induction of mouthpart deformities in Chironomus riparius larvae exposed to 4nnonylphenol. Environ. Pollut. 2000, 111 (2), 241246.

Park, K.; Bang, H.; Park, J.; Kwak, I., Ecotoxicological multilevelevaluation of the effects of fenbendazole exposure to Chironomus riparius larvae. Chemosphere 2009, 77 (3), 359367.

Park, K.; Kwak, I.S., Characterization of heat shock protein 40 and 90 in Chironomus riparius larvae: Effects of di(2ethylhexyl) phthalate exposure on gene expressions and mouthpart deformities. Chemosphere 2008, 74 (1), 8995.

Park, K.; Park, J.; Kim, J.; Kwak, I.S., Biological and molecular responses of Chironomus riparius (Diptera, Chironomidae) to herbicide 2,4D (2,4dichlorophenoxyacetic acid). Comp.

Biochem. Physiol. C Toxicol. Pharmacol. 2010, 151 (4), 439446.

Pellinen, J.; Soimasuo, R., Toxicity of sediments polluted by the pulp and paper industry to a midge (Chironomus riparius Meigen). Sci. Total Environ. 1993, 134 (SUPPL. 2), 12471256.

Segner, H.; Caroll, K.; Fenske, M.; Janssen, C. R.; Maack, G.; Pascoe, D.; Schafers, C.;

Vandenbergh, G. F.; Watts, M.; Wenzel, A., Identification of endocrinedisrupting effects in

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Page 102 aquatic vertebrates and invertebrates: report from the European IDEA project. Ecotoxicol.

Environ. Saf. 2003, 54 (3), 302314.

Song, Z. H., Effects of Pentachlorophenol on Galba pervia, Tubifex sinicus and Chironomus plumousus Larvae. Bull. Environ. Contam. Toxicol. 2007, 79 (3), 278282.

Tomilina, I. I.; Grebenyuk, L. P.; Chuiko, G. M., Toxicological and Teratogenic Assessment of Bottom Sediments from the Rybinsk Reservoir. Inland Water Biol. 2011, 4 (3), 373382.

Tomilina, I. I.; Gremyachikh, V. A.; Grebenyuk, L. P.; Klevleeva, T. R., The effect of zinc oxide nano and microparticles and zinc ions on freshwater organisms of different trophic levels. Inland Water Biol. 2014, 7 (1), 8896.

Vedamanikam, V.; Shazili, N., Observations of mouthpart deformities in the Chironomus larvae exposed to different concentrations of nine heavy metals. Toxicol. Environ. Chem.

2009, 91 (1), 5763.

Vermeulen, A. C.; Dall, P.; Lindegaard, C.; Olliver, F.; Goddeeris, B. R., Exposure of

Chironomus riparius larvae (Diptera) to lead, mercury and ßsitosterol: Effect on mouthpart deformation and moulting. Chemospere 2000, 41, 4555.

Vogt, C.; LangerJaesrich, M.; Elsässer, O.; Schmitt, C.; Van Dongen, S.; Köhler, H. R.;

Oehlmann, J.; Nowak, C., Effects of inbreeding on mouthpart deformities of Chironomus riparius under sublethal pesticide exposure. Environ. Toxicol. Chem. 2013, 32 (2), 423425.

Watts, M. M.; Pascoe, D.; Carroll, K., Exposure to 17 alphaethinylestradiol and bisphenol

Aeffects on larval moulting and mouthpart structure of Chironomus riparius. Ecotoxicol.

Environ. Saf. 2003, 54 (2), 207215.

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Page 103 Wise, R. R.; Pierstorff, C. A.; Nelson, S. L.; Bursek, R. M.; Plude, J. L.; McNello, M.; Hein,

J., Morphological deformities in Chironomus (Chironomidae : Diptera) larvae as indicators of pollution in Lake Winnebago, Wisconsin. J. Great Lakes Res. 2001, 27 (4), 503509.

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Page 104

Table S1 Sign test analysis for treatments analysed in Analysis 1. Positive responses are those where there was a higher deformity frequency in the exposure treatment than control [i.e. ln(response) > 0], zero responses are those where ln(response) = 0, and negative responses are where ln(response) < 0. Studies with < 5 treatments were excluded from the statistical analysis.

Total Contaminant Study Positive responses Non-positive responses Sign test p-value Treatments Total Zero Negative

Copper Kosalwat and Knight62 3 3 0 0 0 NA Janssens de Bisthoven et al.57 63 32 31 0 31 1

Di Veroli et al.23 30 20 10 0 10 0.1

Romanenko et al.63 40 20 20 18 2 1

TOTAL 133 72 61 18 43 0.39

Lead Martinez et al.47 3 2 1 0 1 NA Martinez et al.17 3 3 0 0 0 NA Arambourou et al.61 24 12 12 0 12 1

Janssens de Bisthoven et al.57 72 33 39 0 39 0.56

Di Veroli et al.23 30 18 12 0 12 0.36

Wise et al.64 8 5 3 1 2 0.73

Arambourou et al.24 60 33 27 0 0 0.52

TOTAL 194 101 93 1 65 0.36

Zinc Martinez et al.47 3 2 1 0 1 NA Martinez et al.17 3 3 0 0 0 NA

Di Veroli et al.23 30 20 10 0 0 0.1

TOTAL 30 20 10 0 0 0.1

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Page 105

Figure S1 Funnel plot of experimental treatments for which samples size (number of larvae) was reported. Solid black line represents x = 0 (i.e. “no chemical effect”), red line represents mean log odds ratio (x = 0.46). Dashed lines represent the 95% confidence interval envelope. Egger’s test suggested no funnel plot asymmetry (t = 0.99, df = 444, p = 0.32), and hence no publication bias.

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Page 106 Figure S2 Scatterplot of year of publication of each deformity treatment against log odds ratio. Spearman’s correlation analysis revealed no significant change in effect size over time

(ρ = 0.09, p = 0.052), suggesting no publication bias

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Page 107 Appendix 3 – Supplementary Material to Chapter 3

Page 108 S1 Recovery of larvae (summed across replicate beakers) including statistical analyses, and physicochemical water quality results [mean (SD)] for deformity experiments.

Experiment Condition Treatment n Recovered Unrecovered GLM analysis for n (larvae scored Water quality (replicate larvae larvae recovery differences for deformities) beakers) (including controls) z p Dissolved oxygen pH Electrical (%S) conductivity (µS/cm) Stress Non-stress Control 3 30 0 26 89.84 (1.12) 6.95 (0.25) 191.9 (17.51)

Copper exposure 0.05 3 27 3 -0.246 0.06 23 93.97 (1.55) 6.94 (0.10) 184.27 (15.73) 0.09 3 30 0 26 91.68 (2.95) 6.84 (0.07) 173.3 (1.64) 0.14 3 30 0 24 92.96 (0.92) 7.27 (0.37) 169.93 (3.11) 0.24 3 30 0 26 95.34 (2.22) 7.03 (0.18) 171.83 (6.42) 0.40 3 26 4 22 96.32 (1.40) 6.96 (0.08) 170.33 (2.63)

Imidacloprid 0.17 3 28 2 -0.174 0.86 24 90.95 (2.67) 7.08 (0.60) 183.97 (8.51) exposure 0.28 3 30 0 29 91.05 (5.15) 6.89 (0.20) 181.17 (8.25) 0.47 3 29 1 28 89.54 (3.73) 6.85 (0.17) 186.50 (3.12) 0.78 3 30 0 29 84.21 (11.11) 6.76 (0.08) 192.00 (9.66) 1.30 3 29 1 27 88.18 (1.93) 6.70 (0.08) 188.17 (1.65)

Food limitation 186 3 29 1 0.746 0.46 28 96.86 (1.97) 7.05 (0.08) 184.43 (5.06) 286 3 30 0 30 97.74 (0.94) 7.21 (0.24) 189.10 (1.74) 440 3 30 0 27 95.71 (3.68) 7.28 (0.57) 197.10 (10.85) 676 3 29 1 28 94.65 (4.52) 7.03 (0.18) 203.97 (14.10) 1040 3 30 0 29 92.66 (5.12) 7.13 (0.42) 193.50 (6.64)

Timecourse Onset of 5 (Control) 4 37 3 -1.732 0.08 37 102.32 (2.26) 7.76 (0.44) 254.23 (9.16) pupation 6 4 37 3 37 103 (1.16) 7.46 (0.30) 211.02 (7.84) 7 4 38 2 36 101.00 (0.63) 7.44 (0.37) 205.5 (9.09) 8 4 35 5 34 102.85 (0.63) 7.45 (0.43) 246.78 (10.12) 9 3 24 6 22 102.26 (1.04) 7.44 (0.12) 243.23 (5.36) 10* 4 22 18 0 98.69 (3.29) 7.34 (0.30) 217.55 (15.11) Non-recovery = dead, unrecovered, and < 4th instar living larvae. *not analysed for recovery or deformities due to < 80% recovery

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