Ecotoxicological effects of anthropogenic stressors in subterranean organisms: A review

Journal Pre-proof Ecotoxicological effects of anthropogenic stressors in subterranean organisms: A review Andrea Castaño-Sánchez, Grant C. Hose, Ana Sofia P.S. Reboleira PII:

S0045-6535(19)32662-1

DOI:

https://doi.org/10.1016/j.chemosphere.2019.125422

Reference:

CHEM 125422

To appear in:

ECSN

Received Date: 4 September 2019 Revised Date:

18 November 2019

Accepted Date: 19 November 2019

Please cite this article as: Castaño-Sánchez, A., Hose, G.C., Reboleira, A.S.P.S., Ecotoxicological effects of anthropogenic stressors in subterranean organisms: A review, Chemosphere (2019), doi: https://doi.org/10.1016/j.chemosphere.2019.125422. This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain. © 2019 Published by Elsevier Ltd.

CRediT author statement Andrea Castaño-Sánchez: Conceptualization, Methodology, Visualization, Data Curation, Writing – Original Draft, Investigation, Formal Analysis. Grant Hose: Visualization, Writing – Review & Editing. Ana Sofia Reboleira: Conceptualization, Methodology, Data Curation, Resources, Investigation, Formal Analysis, Funding Acquisition, Project Administration, Supervision.

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Ecotoxicological effects of anthropogenic stressors in subterranean organisms: A review

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Andrea Castaño-Sáncheza; Grant C. Hoseb; Ana Sofia P.S. Reboleiraa *

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a

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2100 Copenhagen, Denmark

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b

Department of Biological Sciences, Macquarie University, NSW 2109 Sydney, Australia

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*

Corresponding author: [email protected]

Natural History Museum of Denmark, University of Copenhagen, Universitetsparken 15,

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Abstract

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How anthropogenic stressors affect biodiversity is a central question in a changing world.

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Subterranean ecosystems and their biodiversity are particularly vulnerable to change, yet,

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these species are frequently neglected in analyses of global biodiversity and assessments of

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ecological status and risk. Are these hidden species affected by anthropogenic stressors? Do

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they survive outside of the current thermal limits of their ecosystems? These and other

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important questions can be addressed with ecotoxicological testing, relating contaminants and

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temperature resistance of species with measured environmental concentrations and climatic

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data. Ecotoxicological knowledge specific to subterranean ecosystems is crucial for

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establishing thresholds for their protection, but such data are both scarce and scattered. Here,

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we review the existing ecotoxicological studies of these impacts to subterranean-adapted

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species. An effort that includes 167 measured endpoints and presents a database containing

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experimentally derived species’ tolerance data for 28 contaminants and temperature, for 46

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terrestrial and groundwater species, including fungi and animals. The lack of standard data

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among the studies is currently the major impediment to evaluate how stressors affect

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subterranean-adapted species and how differently they respond from their relatives at surface. 1

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Improving understanding of ecotoxicological effects on subterranean-adapted species will

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require extensive analysis of physiological responses to a wide range of untested stressors,

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standardization of testing protocols and evaluation of exposures under realistic scenarios.

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Keywords

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Subsurface ecosystems; human impacts; aquifer; ecotoxicological assays; stygofauna;

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troglofauna.

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1. Introduction

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While most of the world's biodiversity is found at or very near the surface, numerous species

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also live deep in the underground (Danielopol et al., 2003; Humphreys, 2006). The

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subterranean domain harbors a peculiar ecosystem, containing narrowly-distributed species

37

with unique morphological and physiological traits, providing an aphotic habitat both in air

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and water filled spaces below the surface (Humphreys, 2006; Dole-Olivier et al., 2009; Culver

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and Pipan, 2019). As ecosystems, they range from shallow subterranean habitats, to deep

40

caves or meso and macrofissures that may be inundated by fresh, brackish or marine waters

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(Culver and Pipan, 2019). As depth below ground increases, the influence of surface

42

processes decreases, and environmental conditions become increasingly constant (stable

43

temperature, high humidity and permanent darkness) (Culver and Pipan, 2019). These stable

44

conditions support diverse communities of highly adapted organisms (mainly invertebrates)

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that have colonized this environment at different times (Gibert and Deharveng, 2002). These

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subterranean ecosystems constitute one of the most poorly known and unprotected natural

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resources of our planet (Pipan et al., 2010).

2

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Access to subterranean environments is often limited to artificial (wells and boreholes) and

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natural points (caves and springs) (Sket, 2018), which is reflected in the relatively limited

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sampling of subterranean environments globally (Schneider and Culver, 2004; Culver et al.,

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2006). Studies of more accessible ecotone habitats such as the aquatic hyporheic zone and the

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terrestrial mesovoid shallow substrate (MSS) have provided valuable insights although the

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extent of knowledge of deeper subterranean habitats is patchy. Indeed, the difficulty of

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accessing subterranean ecosystems combined with the strong geographical biases on species

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distribution have been recently described as the “Racovitzan impediment” (Ficetola et al.,

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2019).

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The infiltration of contaminants into the subsurface can be rapid and difficult to manage,

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making subterranean ecosystems particularly vulnerable to pollution (Marmonier et al., 2013,

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2018). Despite worldwide recognition that subterranean ecosystems are both the most

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important sources of freshwater for human consumption, and critically endangered

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ecosystems, initiatives like the EU Water Framework Directive (EU, 2000), the Groundwater

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Directive (EU-GWD, 2006), the Safe Drinking Water Act of the US EPA (HDR, 2017), the

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guidelines for drinking-water quality from the World Health Organization (WHO, 2011), or

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the Australian Drinking Water Guidelines (NRMMC, 2011), focus only on water quality and

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the need to achieve a good physicochemical status of groundwater, neglecting protection to its

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endemic biodiversity (Reboleira et al., 2013). Subterranean biota play a key role in functions

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that provide ecosystem services, including the maintenance of water quality and the support

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of groundwater dependent ecosystems, such as springs and rivers (Griebler et al., 2014).

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Therefore, the evaluation of the condition of subterranean ecosystems should consider both

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abiotic and biotic components. Furthermore, the fact that terrestrial subterranean ecosystems

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are intimately linked with the groundwater cycle is also neglected and scientific information 3

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concerning the effect of pollution in these ecosystems is needed for their protection.

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Several authors suggested that the evolutionary adaptations of subterranean biota make them

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particularly sensitive to anthropogenic impacts (Griebler et al., 2010; Griebler and Avramov,

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2015; Jiménez Valverde et al., 2017). The primary, global human impact in subterranean

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ecosystems is the extraction of water leading to the depletion of aquifers. Around 50% of the

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world’s population relies on aquifers for drinking water (World Water Assessment

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Programme, 2003), and this is likely to increase under the predicted global warming and

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climate change scenarios (Danielopol et al., 2003; Kløve et al., 2014). In parallel with the

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over extraction of water, subterranean ecosystems are threatened by anthropogenic pollutants

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(Danielopol et al., 2003; Marmonier et al., 2013).

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Ecotoxicological dose-response bioassays are an important line of evidence for environmental

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risk assessments (ERA) because they provide experimental evidence of cause and effect. A

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large number of bioassays have been developed to assess the toxicity of contaminants to

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plants, animals and microbes in terrestrial, marine and freshwater environments (Breitholtz et

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al., 2006). However, the development of bioassays for subterranean organisms has been

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limited. The first reported dose-response toxicity tests using subterranean species was

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conducted in 1976 (Bosnak and Morgan, 1981), and since then, a small number of studies

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have been done using aquatic subterranean organisms. While there is an extensive body of

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literature on bioassays for superficial soil biota (Sochová et al., 2006; Van Gestel, 2012;

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Duarte et al., 2017), there have been no bioassays for assessing the toxicity of pollutants using

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terrestrial subterranean-adapted organisms.

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Here we provide a critical review of the ecotoxicological effects of anthropogenic stressors on

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subterranean organisms, and examine the state of the art of ecotoxicological bioassays with

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subterranean organisms, including those associated with climate change. We synthesized 4

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available data on the ecotoxicological effects known for subterranean species, focusing on the

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type of stressors, studied organisms and experimental conditions, and provide future

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perspectives for ecotoxicology on subterranean ecosystems. This constitutes a fundamental

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starting point for the development of standard methods for estimating the effect of stressors

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and climate change in these hidden ecosystems, definitely contributing to their protection.

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2. The subterranean ecosystem 2.1 Environmental constraints

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Conditions in subterranean ecosystems are typified by the absence of light, high air humidity,

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oligotrophic (low carbon and nutrient) conditions (Schmidt and Hahn, 2012), and stable

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temperatures (relative to surface environments); overall, physical and chemical conditions are

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relatively stable compared to surface systems (Gibert and Deharveng, 2002). This suite of

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conditions are common across all types of subterranean ecosystems (Culver and Pipan, 2019).

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Subterranean ecosystems occur in highly complex and heterogeneous geological settings

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(Mösslacher and Notenboom, 1999) including unconsolidated (e.g. alluvial sediments), and

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consolidated rocks (e.g. karst). The connectivity within and between subterranean habitats,

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and between them and the surface is determined by the connectivity among voids of the

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matrix. Habitats can thus range from those that are highly and homogeneously porous and

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well connected to the surface and adjoining habitats, to those that are highly complex and

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fragmented with isolated pockets, and limited vertical and lateral connectivity. This

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architecture of spatial partitioning and high isolation presents a pattern similar to island

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biogeography, and has resulted in strong spatial heterogeneity and patchiness in species

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distributions (Dole-Olivier et al., 2009; Malard et al., 2009).

5

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The size of voids within the geological matrix, and the environmental conditions have been

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primary selection forces for the species that have successfully colonized subterranean

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ecosystems (e.g. Korbel and Hose, 2015). The size of the voids in the subterranean matrix

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limit the size of biota that can inhabit it (Korbel et al., 2019). As a consequence, vertebrates

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are relatively uncommon and biota is dominated by microbes (such as bacteria, fungi and

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protozoa) and invertebrates (macro and meiofauna) (Gibert and Deharveng, 2002; Griebler

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and Lueders, 2009; Pipan and Culver, 2017).

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2.2 Simplified trophic chain

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The subterranean food web is often strongly truncated, basally and apically, due to the

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absence of light and consequently of photosynthetic producers, and potentially low abundance

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of strict predators (Fig. 1) (Gibert and Deharveng, 2002). With the general absence of

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vertebrates, invertebrates are often the highest trophic level. In the absence of photosynthetic

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organisms, subterranean ecosystems are largely reliant on carbon infiltrating from the surface,

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with few exceptions where there is significant primary production from chemolithotrophic

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bacteria (Sarbu et al., 1996; Kumaresan et al., 2014; Brankovits et al., 2017). As a

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consequence, carbon is often limited and microbial diversity and activity are typically low

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(Griebler and Lueders, 2009). Microbial diversity found in the subsurface is represented by

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Archaea, Bacteria, Protozoa and Fungi, which remain largely unknown (Griebler and Lueders,

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2009). The scarcity of carbon and nutrients constrains population densities (biomass) and has

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forced species to adapt and cope with a low energy environment by decreasing metabolic

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rates and enhancing food detection through chemoreception and well-developed sensory

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appendages (Di Lorenzo et al., 2015a). Also, functional redundancy, i.e., several species that

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fulfill the same role (food web) within the ecosystem might occur (Gibert and Deharveng, 6

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2002; Korbel and Hose, 2011; Saccó et al., 2019a), however specific feeding traits have been

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observed in the same trophic level, even within the same genus (Arnedo et al., 2008; Hutchins

145

et al., 2014; Francois et al., 2015).

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2.3 Ecological classification of biota

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The environmental constraints of the subterranean domain provide strong selective pressures

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and have led to similar morphological, physiological and life-history traits across independent

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evolutionary lineages of organisms (Notenboom et al., 1994; Hancock et al., 2005;

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Humphreys, 2006). These specialized traits have been named troglomorphisms or

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“troglomorphic syndrome” (Christiansen, 2012). The ecological classification of subterranean

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species is based on their degree of dependence on the subterranean ecosystems (White et al.,

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2019). Species that complete their whole life-cycle in the subterranean environment and

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typically exhibit typical morpho-physiological traits to the subterranean environment (e.g.,

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lack of pigment and ocular regression, body and appendages elongation, hypertrophy of

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sensory organs, low metabolic rate and prolonged life span with low reproductive output) are

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referred to as troglobiont (terrestrial) or stygobiont (aquatic). The suffix -phile

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(troglophile/stygophile) is used for those species with affinities to subterranean ecosystems,

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but can also complete their life-cycle at the surface. Finally, the suffix -xene

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(trogloxene/stygoxene) is used to describe species that are present sporadically in

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subterranean ecosystems but are unable to establish a subterranean population; they are

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considered occasional inhabitants (Sket, 2008).

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3. Anthropogenic stressors on the subterranean environment

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Anthropogenic pressures have surpassed natural forces as drivers of change in Earth’s

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environment (Steffen et al., 2007). Subsoil and groundwater contamination, groundwater

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pumping and mining activities threaten subterranean ecosystems and the provision of

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ecosystem services (Danielopol et al., 2003; Larned, 2012) (Fig. 2). Moreover, the changes to

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the atmosphere and surface environments associated with climate change (warming effect)

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will result in changes to the subsurface quality (Dijck et al., 2006) and its annual thermal

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regime, but the lack of knowledge makes impossible to determine the magnitude and direction

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to climate related trends for subterranean ecosystems (Kundzewicz and Döll, 2009).

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The main cause of subterranean pollution is by infiltration from surface due to industry, urban

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discharge (urban water disposals), transport, agriculture and farming (Marmonier et al., 2013).

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Contaminants frequently detected in subterranean ecosystems include metals, pesticides,

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fertilizers, emerging contaminants, and volatile organic compounds (VOCs) (Fig. 3, Table

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S.1). Measured environmental concentrations (MECs) at which these contaminants are

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measured in groundwater can be found in databases such as EEA (2019) and the National

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Water Quality Monitoring (2019), among others. However, similar information for sediment

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contamination in the subterranean terrestrial and aquatic compartments are still missing. The

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impact of contaminants depends on the source, physicochemical characteristics, mobility and

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behavior in soil or the aqueous environment and associated hazards (Stuart et al., 2012). In the

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following sections we examine the current extent of common contaminants in subterranean

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ecosystems.

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3.1 Metals

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Metal (and metalloid) contamination of subterranean and surface environments has received

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more attention than other contaminants because of the long history of metals use (Momodu 8

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and Anyakora, 2010; Kumar et al., 2012) (Fig. 3A). The toxicity, biological activity,

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bioavailability, solubility, mobility, reactivity and elimination of a metal can vary depending

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on its speciation form (Santos et al., 2002), which itself depends on the physicochemical

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characteristics of the environment (such as pH, redox potential, temperature, complexation

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with other dissolved constituents, sorption and ion-exchange capacity of the geological

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materials, and organic matter content) (Evanko and Dzombak, 1997; Hashim et al., 2011).

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Accordingly, contamination by metals in karst areas needs special consideration, since the

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alkalinity (> 7 pH) and the carbonates may remove metals from the dissolved phase (Vesper,

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2019). The formation of metal-carbonate precipitates and secondary minerals reduces the

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metal solubility and transport. The resulting chemical species will have different toxicity and

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bioavailability (White et al., 2019).

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Anthropogenic sources of metals in subterranean environments include point sources such as

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mining activities (in which contamination may continue long after mining activities have

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ceased (Hutton and Symon, 1986; Nriagu, 1989; Duruibe et al., 2007)), biosolid and sludge

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applications to land, and waste water discharges, as well as diffuse sources such as urban and

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agricultural runoff (Evanko and Dzombak, 1997). Diffuse metal contamination is a significant

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problem in some agricultural areas (Vodela et al., 1997; Wongsasuluk et al., 2014) since

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metals are common constituents of pesticides and fertilizers (Santos et al., 2002).

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3.2 Pesticides

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Pesticides detected in subterranean ecosystems (Fig. 3B) include synthetic chemicals and

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natural substances that can be classified according to their specific biological activity on

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target species (Van der Hoff and Van Zoonen, 1999; Ahmed, 2001). As a result, this group

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includes chemicals with a wide variety of uses, properties and potential impacts on the 9

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environment (Stuart et al., 2012). Over 500 compounds are currently registered worldwide as

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pesticides, or metabolites of pesticides (Van der Hoff and Van Zoonen, 1999; Ahmed, 2001;

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Loos et al., 2010).

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Pesticides are considered a diffuse source of pollution for subterranean environments since

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they are typically applied over broad geographical areas (Stuart et al., 2012). Most pesticides

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infiltrate by percolating through the soil profile, or in the case of caves, mobilization of

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contaminated soil from the surface. The mobility of a pesticide in the environment depends

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heavily on its chemical properties, particularly its solubility in water and its affinity for

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binding to soil and organic matter (Arias-Estévez et al., 2008). Degradation of pesticides is

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governed by both abiotic and biotic processes, with the degradation products also biologically

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active and often equally or more toxic than the parent compound (Arias-Estévez et al., 2008;

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Stuart et al., 2012).

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Studies have shown that shallow groundwater is more vulnerable to contamination from

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pesticides than deep groundwater, because of its proximity to the surface (Haarstad and

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Ludvigsen, 2007; Close and Skinner, 2012; Toccalino et al., 2014). Moreover, Stuart et al.

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(2012) emphasize the importance of the rapid transport routes (like drains, sinkholes and

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fractures), which are especially common in karst areas, that by-pass the main natural pesticide

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attenuation areas in the vadose zone (Jiménez-Sánchez et al., 2008).

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3.3 Fertilizers

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The largest source of nutrients in groundwater is nitrogen and phosphorous – based fertilizers

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used in agriculture (Fig. 3C). These can be synthetic chemical fertilizers or organic-waste

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fertilizers (e.g. compost and blood or bone meal) which contain essential and nonessential

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elements as metals (Khan et al., 2018). Both nitrogen and phosphorous are essential macro-

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nutrients for plant growth and are thus critical for primary production in the biosphere; 10

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however, when present at concentrations exceeding natural background levels they can cause

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eutrophication problems (Gruber and Galloway, 2008).

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Most of the world’s phosphate fertilizers come from phosphate rocks which also frequently

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contain hazardous elements such as Cd, Pb, Hg, U, Cr and As. The application of phosphate

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fertilizers can promote the contamination of agricultural soils by introducing these hazardous

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metals and enhancing the mobility of metals already in the soils (Campos, 2002; Dissanayake

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and Chandrajith, 2009; Khan et al., 2018).

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The production of ammonia by the reaction of nitrogen and hydrogen, known as the Haber-

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Bosh process, has enabled humans to approximately double the input of available nitrogen to

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the Earth’s land surface, and greatly increase food production through the use of nitrogenous

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fertilizers. However, this increase in nitrogen in a readily bioavailable form has also created

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environmental issues (Galloway et al., 2008; Gruber and Galloway, 2008). Nitrogen is a

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particular threat to groundwaters because of its high solubility in its various forms.

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Background concentrations of nitrate in groundwater are typically below 2 mg/L,

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consequently concentrations over 3 mg/L are considered indicative of anthropogenic inputs

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(Burkart and Stoner, 2008).

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Factors such as climate, soil properties, groundwater fluxes and management practices

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influence the fertilizer load entering subterranean ecosystems (Wang et al., 2010). However,

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the relationships between cropping systems, fertilizer demand, soil nitrate and phosphate

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accumulation and groundwater fertilizers concentrations are still not fully understood (Ju et

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al., 2006). Other nutrient sources in groundwater, include leakage from sewage systems and

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septic tanks, industrial spillages, landfills, river or channel infiltration, and fertilizers used in

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gardens (Wakida and Lerner, 2005). As for pesticides, shallow water-table aquifers and karst

11

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systems are more vulnerable to nutrient contamination (Bijay-Singh et al., 1995; Boyer and

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Pasquarell, 1995; Jiménez-Sánchez et al., 2008).

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3.4 Volatile organic compounds (VOCs)

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Volatile organic compounds (VOCs) are among the most frequently detected contaminants in

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groundwaters (Squillace et al., 2002; EEA, 2007), usually at concentrations in the order of

267

µg/L (Fig. 3D). They are carbon-based compounds that evaporate readily at normal

268

temperatures (Fan et al., 2009) and include some oxygenated compounds, such as ethers or

269

aromatic hydrocarbons from mineral oil (e.g. benzene, toluene), and chlorinated aliphatic

270

hydrocarbons (Rivett et al., 2011).

271

Due to their volatility, VOCs enter the atmosphere in large quantities from anthropogenic

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(e.g. biomass burning, oil and gas industry emissions) and biogenic sources (e.g. natural

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wetland emission and domestic ruminants) (Atkinson, 2000; Fan et al., 2009). The

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atmospheric deposition is the main non-point/diffusive source of VOCs in subterranean

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ecosystems whereas point sources are primarily industrial or waste disposal sites where the

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use and production or inadequate storage or disposal contribute VOCs to the soil or

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groundwater (Pankow et al., 1997; Fan et al., 2009; Stephenson et al., 2013). Once in the soil,

278

VOCs can undergo diffusive transport through the unsaturated zone as a gas-phase or

279

infiltrate into the groundwater. In groundwater, VOCs may act as a non-aqueous phase liquid

280

(NAPL) due to their limited solubility in water, allowing them to remain in a separate phase.

281

Aromatic hydrocarbons often form light non-aqueous-phase liquids (LNAPL) and chlorinated

282

aliphatic hydrocarbons form denser non-aqueous-phase liquids (DNAPLs) (White et al., 2019;

283

Rivett et al., 2011). Once in groundwater, the DNAPLs, dissolve very slowly creating plumes

284

that persist over decades (Loop and White, 2001; Ellis and Rivett, 2007). 12

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Field studies have shown changes in subterranean communities in response to VOC

286

contamination (e.g. Stephenson et al., 2013), but few laboratory-based studies have been

287

conducted. Due to their physical and (bio) chemical features such as volatility, limited

288

solubility in water, biodegradability, as well as sorption and bioaccumulation potential, VOCs

289

“present difficulties in the execution and interpretation of standard aquatic toxicity tests”

290

(Avramov et al., 2013).

291 292

3.5 Emerging contaminants

293

Emerging contaminants are substances that may have been present for a long time in the

294

environment, but whose presence is now being detected continuously at subtoxic

295

concentrations in groundwater (Fig. 3E) and can cause unexpected consequences in humans

296

and ecosystems (Daughton, 2005). Emerging contaminants include pharmaceuticals, personal

297

care

298

perfluorooctanesulfonic acid compounds (PFOS) (see Barceló, 2003).

299

The behavior of emerging contaminants in the subsoil, their solubility or ability to travel

300

through the aqueous environment, and persistence in the subsurface environment is generally

301

poorly known (Stuart et al., 2012). They occur mostly as mixtures, meaning that organisms

302

are usually exposed to complex multi-component mixtures over their whole life cycle

303

(Barceló, 2003).

products,

UV

filters/sunscreen,

food

additives,

flame/fire

retardants

and

304 305

3.6 Temperature

306

Subterranean-adapted organisms are typically exposed to small thermal fluctuations in their

307

environment. They are considered mostly ectothermic with a thermal tolerance is within the

308

range of conditions experienced in their natural habitat (Addo-Bediako et al., 2000), however 13

309

some stygobionts exhibit eurythermal characteristics (Issartel et al., 2005). Increasing

310

temperatures associated with climate change may affect the quality of subterranean

311

environments in two ways (Kløve et al., 2014). Firstly, expected changes in recharge rates

312

may affect the mobilization of compounds in the unsaturated zone, and, secondly, changes to

313

thermal regimes have important implications for the rates of temperature-dependent reactions

314

that directly affect many biogeochemical processes (e.g. redox reactions linked to nitrogen

315

and carbon cycles, and the speciation and mobilization of contaminants). These effects all

316

pose a risk to subterranean biodiversity.

317 318

4. Effects of anthropogenic stressors on subterranean organisms

319

Ecotoxicological testing is a useful tool for assessing the tolerance of organisms to

320

contaminants that threaten subterranean environments (Avramov et al., 2013). Bioassays are

321

based on dose/exposure-response assessment, which quantify the relationship between dose or

322

degree of exposure to a substance and the toxic effect or response (EC, 2003). They span a

323

continuum of biological complexity and ecological realism, from small, rapid and highly

324

reproducible tests on populations under laboratory conditions to long-term, large scale,

325

complex field-based community studies (Breitholtz et al., 2006). Bioassays vary in duration,

326

typically considered as either: a) acute assays – short-term exposure test, used to observed the

327

effects of toxicants under a single event exposure, where lethality is the most common

328

endpoint; or b) chronic assays – longer-term tests, with (usually) lower and continuous

329

contaminant exposure, mostly used to estimate sub-lethal endpoints related to the organism’s

330

life span (population effect) (Van Leeuwen et al., 1996).

331

The toxicity values reported in ecotoxicological bioassays, together with the environmental

332

contaminant exposure assessment are key lines of evidence in ERA (EC, 2003). Importantly, 14

333

the toxicity data used in ERA should reflect the species likely to be affected (EC, 2003;

334

Breitholtz et al., 2006; Kulkarni et al., 2013). The current lack of such data is a critical

335

knowledge gap for determining ecological thresholds values for subterranean ecosystems (Di

336

Lorenzo et al., 2015b).

337

Ecotoxicity data reported here for subterranean fauna were obtained from the available

338

literature using: i) Google Scholar, Web of Science and Scopus, with the following strings

339

applied to the topic field: “stygo*”, “troglo*”, “subterranean”, “groundwater”, “thermal

340

tolerance” “sensitivity” and “ecotox*”; and ii) EXPLORE of the US EPA ECOTOX database

341

(https://cfpub.epa.gov/ecotox/explore.cfm). A detailed literature survey was done to review all

342

ecotoxicological data available for subterranean organisms, targeting specifically: 1) organism

343

information (species and ecological classification); 2) organism origin and manipulation

344

before the test start (specimens origin and stage, acclimation to the laboratory condition and

345

test condition); 3) contaminant information (name, class); 3) type of test performed (test

346

media, temperature, aeration, light, feeding and duration) and; 4) evaluation of the estimated

347

effect results (end point and toxicity measured). Data are summarized in Table S.2 for

348

contaminants and Table S.3 for temperature (supplementary material).

349 350

4.1 Contaminants tested on subterranean organisms

351

Subterranean organisms have been tested mostly in acute assays, under conditions of short,

352

often high-concentration exposures and starvation conditions with mortality as end points

353

(Table S.2). Among all studies there was considerable variability in the protocols used. Test

354

organisms were field-collected (rather than laboratory cultures), mostly from Europe, but also

355

from North Africa, western USA and southeastern Australia (Fig. 4). Pre-test acclimation

356

periods ranged from a few hours to eight months and only in some cases the organisms were 15

357

under starvation. The life stage, size or sex of the organism tested was not always recorded

358

and the number of tested organisms per concentration varied between 6 and 30. The medium

359

used in some cases was aerated or renewed and varied from environmental to artificial water,

360

(except for fungi tested in agar) and the temperature ranged from 10.5 to 22 °C. Some studies

361

added food and most were conducted in darkness. The end point measured was mortality

362

(with the exception of growth inhibition in a fungus) expressed most commonly as the lethal

363

(LC50) or effective (EC50) concentrations for half of the population, or as lethal concentration

364

values which affect a specified proportion (e.g. 10, 25, 100%) of the population. The exposure

365

periods (24 h to 28 d) were always shorter than the organism’s life cycle (Table S.2).

366

The majority of toxicity studies with subterranean organisms tested the effects of metals,

367

followed in number by pesticides, fertilizers, chlorine and volatile organic compounds. In all

368

experiments, the contaminants were tested individually (Table S.2).

369

To date most species tested have been groundwater (stygobiont and stygophile) crustaceans:

370

the copepods Bryocamptus echinatus, Diacyclops belgicus, Parastenocaris germanica, two

371

undescribed species of cyclopoids “Cyclopioda_sp1 and Cyclopoida_sp2” and one

372

harpaticoid “Harpaticoida_sp1”; the isopods Caecidotea bicrenata, C. stygia, Proasellus

373

assaforensis, P. cavaticus, P. lusitanicus, and Typhlocirolana haouzensis; the amphipods

374

Metacrangonyx spinicaudatus, Niphargus aquilex and N. rhenorhodanesis; and the decapod

375

Orconectes australis. Non-crustacean tested taxa include the stygobiont Trichodrilus tenuis

376

(annelid worm) and the stygophile fungi Penicillium sp. and Rhodotorula minuta (Table S.2).

377

The sensitivity of any terrestrial subterranean species to toxicants remains unknown.

378 379

4.1.1 Metals

16

380

Metals are the most tested compounds, representing 69% of available data. Acute toxicity has

381

only been tested in subterranean organisms for seven metals and metalloids (As, Cd, Cr, Cu,

382

Ni, Pb and Zn) (Fig. 5). The various studies reported either of two end points: lethal

383

concentration measured at different percentage (e.g., IC25, LC10/LC50) or toxicity data for

384

more than one period (e.g., test duration ranged from 24 h to 28 d) (Table S.2.). The

385

sensitivity responses of the different tested taxa to each metal are variable and no taxon was

386

consistently more or less sensitive, nor was any metal (Fig. 5, Table S.2).

387

The results reflect large variability in the methodological test conditions (see Table S.2),

388

especially regarding the end points and length of the experiment, which confound

389

comparisons of responses between taxa. Bearing this in mind, some patterns in sensitivity are

390

still evident within and between taxonomic groups. For example, the response of stygobiont

391

copepods to Cr and Zn was similar among the three species tested, while stygobiont isopods

392

showed a wide range of sensitivity to Cr, Cu and Zn among the five species tested. There was

393

more than one order of magnitude variation in sensitivity to Cu and Cr at the genus level in

394

Proasellus spp (48-h EC50 = 1.12 mg Cr/L and 6.21 mg Cu/L for P. lusitanicus, and 48-h

395

EC50 = 17.9 mg Cr/L and > 52 mg Cu/L for P. assaforensis), and within the genus

396

Caecidotea, C. bicrenata (96-h LC50 = 2.20 mg Cd/L) was considerably more tolerant to Cd

397

than C. stygia (96-h LC50 = 0.29 mg Cd/L). However, C. bicrenata and C. stygia showed

398

similar sensitivity to Cu (96-h LC50 = 2.2 and 2.3 mg Cu/L, respectively).

399

A wide range of sensitivity was observed across groundwater species, especially for As (from

400

24-h EC50 = 4.4 mg As/L for Rhodotorula minuta to 24-h IC25 = 330 mg As/L for Penicillium

401

sp.). Sensitivities to Cd (from 96-h LC50 = 0.29 mg Cd/L for Caecidotea stygia, to 48-h LC50

402

= 150 mg Cd/L for Typhlocirolana haouzensis), and Zn (from 48-h LC50 = 0.45 mg Zn/L for

403

Metacrangonyx spinicaudatus to 96-h LC50 = 118-180 mg Zn/L for Niphargus aquilex) were 17

404

similarly variable. In contrast, the range of response to Cu and Cr were relatively narrow

405

(from 24-h IC25 = 0.02 mg Cu/L and 0.13 mg Cr/L for Penicillium sp, to 48-h EC50 = 17.9 mg

406

Cr/L and > 52 mg Cu/L for P. assaforensis). The effects of Pb and Ni have been even scarcely

407

tested (Fig. 5, Table S.2).

408

In the absence of data from chronic exposure, some authors expanded the exposure time using

409

the acute test conditions until 28 d, to account for the potential slower response in animals

410

with low metabolic rates (e.g., Canivet et al., 2001; Hose et al., 2016). The major weakness of

411

this approach lies in the fact that extending length of the acute test, will likely affect the

412

results, because it is performed under starvation conditions.

413 414

4.1.2 Pesticides

415

Assays with pesticides represent 16% of all assays using subterranean organisms. Ten

416

different pesticides have been tested (α-endosulfan, 3,4-dichlorophenol, aldicarb, ArianeTM,

417

chlorpyrifos, desethylatrazine, Imazamox, penthaclorophenol, S-metolachlor and thiram) in

418

three species of copepods and one amphipod. The range of the sensitivity observed among the

419

pesticides is broad (e.g. 96-h LC50 = 0.003 mg thiram/L for P. germanica, up to 96-h LC50 =

420

199 mg Imazamox/L for Diacyclops belgicus). Aldicarb was the only pesticide tested in more

421

than one species and the sensitivity was similar for both (B. echinatus 96-h LC50 = 2.71 mg/L,

422

and P. germanica 96-h LC50 = 2.99 mg/L) (Table S.2).

423 424

4.1.3 Fertilizers

425

Ammonium chloride (NH4Cl), ionized ammonia (NH4⁺), and urea (CH4N2O) are the

426

fertilizers that have been tested. Ammonium chloride was more toxic to the amphipod M.

427

spinicaudatus (48-h LC50 = 1.4 mg NH4/L) than to the isopod T. haouzensis (48-h LC50 = 96 18

428

mg NH4/L). Similar LC50 values were reported for the stygophile copepods B. echinatus (96-h

429

LC50 = 14.61 mg/L) and D. belgicus (96-h LC50 = 16.33 mg/L) exposed to ionized ammonia.

430

Di Lorenzo et al. (2014) suggest that the LC50 value for urea (96-h LC50 = 3.14 g/L) was the

431

least toxic of the fertilizers that have been tested (Table S.2).

432 433

4.1.4 Total residuals chlorine

434

Total residual chlorine was tested with two different species. The isopod C. bicrenata (96-h

435

LC50 = 0.11 mg/L) and the decapod Orconectes australis australis (24-h LC50 = 2.7-3.39

436

mg/L) (Table S.2).

437 438

4.1.5 VOCs

439

The toxicity of toluene was studied using the amphipod Niphargus inopinatus. The 96-h LC50

440

values (49-63.9 mg/L) were in the same range of magnitude than the LC50 measured after 13-

441

18 d (39-6-47.8 mg/L) (Table S.2).

442 443

4.1.6 Mixtures

444

Three studies have reported the effect of a toxicant mixture on subterranean organisms.

445

Canivet and Gibert (2002) tested the amphipod N. rhenorhodanensis in a complex mixture of

446

metals using maturated lead secondary smelting slags, where the dilutions were obtained

447

mixing fractions of slags from different percolation stages with river water. Di Marzio et al.

448

(2018) exposed copepods to mixtures of the pesticide Imamox and NH4⁺, and the toxicity of

449

the mixture was synergistic (96-h EC50 = 2.71 mg/L), i.e., it was more toxic than the exposure

450

to the two compounds individually (Imamox 96-h LC50 = 199.23 mg/L; NH4⁺ 96-h LC50

451

16.33 mg/L). No interaction was found between S-metolachlor and desethylatrazine on N. 19

=

452

rhenorhodanensis, with high survival recorded in the highest concentration tested (Maazouzi

453

et al., 2016) (Table S.2).

454 455

4.2 Temperature tested on subterranean organisms

456

Thermal tolerance was studied in terrestrial and aquatic subterranean organisms, however

457

there was considerable variability in the methods used (Table S.3). The end points measured

458

were mainly mortality expressed in upper and lower thermal tolerance, while few included

459

sublethal studies on oxygen consumption as a function of temperature and heat shock proteins

460

synthetization (Table S.3). Overall, the experimental approaches used either fixed

461

temperatures or ramping (increasing or decreasing) temperatures, with between 1 h to three

462

months exposure, or until specimens’ death. Results were expressed in days, temperatures,

463

standard respiration rates or survival percentage for a tested temperature (Table S.3).

464

Four groundwater species have been tested for thermal tolerance. The copepod D. belgicus

465

did not express a change in metabolic activity at temperatures up to 1.1 ºC above its actual

466

maximum environmental temperature (Di Lorenzo and Galassi, 2017). The amphipod N.

467

rhenorhodanensis survived within the range of -2 to 28 ºC. They also accumulated

468

cryoprotective molecules and maintained locomotory activity and aerobic metabolism even at

469

-2 ºC (Issartel et al., 2005). The isopod Proasellus valdensi had wide thermal tolerance, with

470

90% survival between 2 to 16 ºC (Mermillod-Blondin et al., 2013). Only the isopods P. “n.

471

sp. 1” and P. “n. sp. 2” were stenothermic, i.e., surviving within a very narrow thermal niche

472

breadth (Mermillod-Blondin et al., 2013).

473

Thermal tolerance has been tested in 17 terrestrial species (Table S.3). Two beetles from the

474

genus Neobathyscia (N. mancinii and N. pasai) and four beetle imagoes: Trapezodirus

475

arcticollis, Macharoscelis infernus, and two species of Troglocharinus (T. fonti and T. ferreri) 20

476

were tested. The beetles tolerated temperatures from 1 to 20 ºC, but none of the four species

477

survived more than 24 h at 25 ºC. Beetles were exposed to gradually increasing temperatures

478

for less than 24 h, and demonstrated a broad range temperature tolerance. Troglocharinus

479

fonti and T. ferreri had a lower thermal limit (LTL) of -2.5 °C and an upper thermal limit

480

(UTL) of  50.7 °C, while the LTL50 was -11.72°C for N. mancinii and -16.96 °C for N. pasai,

481

and their UTL50 was 28 °C. Sublethal assays reported the synthesis of heat shock proteins

482

(HSPs) by N. mancinii and N. pasai at a temperature close to their UTL50. However, this

483

seems to be unspecific for subterranean species, as the synthesis of these HTPs not always

484

occur in stenothermic organisms, e.g., artic sea organisms (Bernabò et al., 2011). The viability

485

of one lineage of subterranean beetles to cope with climate change was also studied using

486

bioclimatic models. Around 60% of the species were predicted to have suitable conditions

487

(within their physiological tolerance range) under different predicted scenarios to 2080, but

488

most of the species were predicted to be exposed to temperatures and rates of change that they

489

have never experienced through their evolutionary history (Sánchez-Fernández et al., 2016).

490

Thirteen Collembola belonging to three categories of adaptation to subterranean ecosystems

491

(nine troglophiles and four troglobionts) from the western Carpathians, were exposed to

492

gradually increasing and decreasing temperatures (ramping rate of 0.15 ºC/min) before and

493

after being exposed for 1 h to the test temperature (7 to 14 temperatures in a step of 0.6 ºC). In

494

all species, the UTL50 (expressed as heat 50% lethal dose) values were higher than the annual

495

maximum temperature in the respective cave, in which the wider thermal tolerant species was

496

Hypogastura crassaegranulata (from a cold lethal dose, LDc50 = -6.9 °C to a heat lethal dose,

497

LDh50 =  36.6 °C), the subtroglophile Tetrodontophora bielanensis was the most cold-sensitive

498

species (LDc50 = -4.4 °C) and troglobiont Pseudosinella paclti the most heat-sensitive species

21

499

(LDh50 =  31.3 °C), results which outline that cold resistance in Collembola is negatively

500

correlated with species body length. There were no significant differences in upper thermal

501

values between different ecological groups. However, it appears that species belonging to

502

ecological groups less associated with the stable cave environments (trogloxenes,

503

troglophiles) had a wider range of temperature tolerance than the more cave-adapted species

504

(troglophiles or troglobionts), and all troglobionts from deeper caves, were markedly heat-

505

sensitive (Raschmanova et al., 2018).

506 507

5. Discussion

508

Data on the responses of subterranean species to anthropogenic stressors come mostly from

509

measuring the mortality of field-collected organisms. However, experimental conditions

510

varied widely between studies which complicates comparisons of results. The selection of test

511

organisms highlights a strong geographic bias, and the rather ad hoc nature of current research

512

in this field, where the primary criterion for collection site and test taxa was the availability of

513

specimens, the accessibility of sampling points and proximity to the laboratory. The effect of

514

contaminants has not been tested with terrestrial fauna, while the effect of temperature was

515

observed with both terrestrial and aquatic subterranean organisms. The great majority of taxa

516

tested were crustaceans, which corresponds with them being often the most frequently

517

encountered and abundant taxa in groundwater (Sket, 2018). There were more data on the

518

ecotoxicological effects of metals to subterranean organisms than for other toxicants, with test

519

endpoints mostly related to organism sensitivity, however some data on bioaccumulation were

520

also reported for the troglophile millipede Apfelbeckia insculpta (Vranković et al., 2017), the

521

stygobiont amphipods N. rhenorhodanesis (Plénet, 1999; Canivet et al., 2001; Canivet and

522

Gibert, 2002) and N. montelianus (Krupa and Guidolin, 2003), the cave fish Clarias 22

523

gariepinus (Du Preez and Wepener, 2016), and the cave salamander Proteus anguinus

524

(Pezdirc et al., 2011). According to EU (2003), bioaccumulation may be used for secondary

525

poisoning assessment to understand the impacts in the trophic chain, which may be

526

particularly pertinent for subterranean ecosystems, where the trophic chains are simplified.

527

Currently, there is no indication that subterranean organisms are more or less sensitive than

528

surface organisms to pollutants. This is a major knowledge gap that limits environmental risk

529

assessments and the setting of specific thresholds for the protection of subterranean

530

ecosystems. To address this central question, focus should be placed in studying the

531

comparative responses of species of the same genus that have surface closely related species,

532

(e.g., across European asellids), in a similar way that was done for surface vs. subterranean

533

populations of the same species (Jemec et al., 2017).

534

Following the initial research focus on metal toxicity, more recent ecotoxicological

535

approaches with subterranean species have followed general trends of research in surface

536

ecosystems, which have focused increasingly on mixtures, pesticides, VOCs, and temperature

537

increases, where sublethal end points have also been used (Guillén et al., 2012; Gavrilescu et

538

al., 2015). However, the progress of such research in subterranean ecosystems is limited by:

539

a) the lack of knowledge of the concentrations of emerging contaminants and fate of these

540

contaminants in subterranean environments (Stuart et al., 2012), especially in the sediment of

541

both aquifer and terrestrial compartment; b) the limited knowledge of subterranean biota and

542

ecosystems structure; and c) the lack of standardized protocols for subterranean ecotoxicology

543

assays (Gibert and Deharveng, 2002; Danielopol et al., 2004; Hose, 2005; Griebler et al.,

544

2010; Reboleira et al. 2013; Di Lorenzo et al., 2014, 2019). As a consequence, there are

545

relatively few toxicity data available compared to the large amount available for surface

546

ecosystems. 23

547

The three main constraints on the use of subterranean organisms for estimating the effects of

548

contaminants in the subsurface are:

549

i.

No standard criteria for choosing species and selection of model/target species. The

550

well-known “short-range” endemic distribution among all taxonomic groups of

551

subsurface-adapted organisms, leads to an absence of widely distributed species that

552

can be used as model/target species for ecotoxicology studies. Furthermore, the

553

subterranean ecosystem is difficult to access and the populations are usually small

554

(Danielopol et al., 2003; Hose, 2005; Gibert et al., 2009; Daam et al., 2010;

555

Reboleira et al., 2013).

556

ii.

The obstacles to developing stable laboratory cultures with subterranean organisms.

557

The limited knowledge of life history and ecological requirements of subterranean

558

fauna and microbes, and typically low reproductive rates, remain an impediment to

559

their successful culturing in the laboratory. Other traits of subterranean organisms,

560

such as slow metabolism, long life cycles, and low thermal tolerance also complicate

561

the rearing of organisms under laboratory conditions (Thulin and Hahn, 2008;

562

Avramov et al., 2013).

563

iii.

Taxonomic

impediments.

Due

to

morphological

convergence,

identifying

564

subterranean species is challenging and time consuming, requiring proper taxonomic

565

training (Ficetola et al., 2019). The absence of a commercial stock of subterranean

566

species requires the use of field collected animals for assays, in which the species

567

identification based on morphology is not always easy to be achieved (Curini-Galletti

568

et al., 2012; Fonseca et al., 2014; Daam et al., 2010; Hose et al., 2016).

569

Despite these challenges, the management of anthropogenic pressures on subterranean

570

ecosystems requires comprehensive multidisciplinary policy framework that recognizes and 24

571

protects both human and ecological values including ecosystems services (Artigas et al.,

572

2012; Saccò et al., 2019b). Ecotoxicological testing of subterranean organisms remains

573

critical for the protection of subterranean ecosystems, and the standardization of testing

574

protocols, including terrestrial species, is urgently needed in order to progress this research

575

field (Di Lorenzo et al., 2019). Ecotoxicological data are a critical line of evidence in ERA,

576

yet testing approaches lag behind other lines of evidence, such as field biomonitoring data, for

577

which indicators and protocols are established (Danielopol et al., 2004; Danielopol and

578

Griebler, 2008; Griebler et al., 2010; Korbel and Hose, 2011, 2017) or the use of sentinel

579

organisms in situ for real life exposure assessment, which is a valuable first step toward

580

including ecotoxicological data in biological indices (Marmonier et al., 2013, 2018). Even the

581

legislation is ahead of the science; the need for robust scientific data specifically for

582

subterranean ecosystems is clearly articulated in some legislations and environmental

583

agencies. For example, the EU-GWD states that “Research should be conducted in order to

584

provide better criteria for ensuring groundwater ecosystem quality and protection. Where

585

necessary, the findings obtained should be taken into account when implementing or revising

586

this Directive” (EU-GWD, 2006).

587

There is a growing appreciation globally of the need to protect subterranean ecosystems

588

(Mammola et al., 2019) through conservation planning and risk assessments that are specific

589

to subterranean ecosystems (Boulton, 2005, 2009; Linke et al., 2019; Di Lorenzo et al., 2019).

590

Subterranean ecosystems and their fauna need urgent protection from anthropogenic activities

591

because:

592

i.

Subterranean fauna have a high risk of extinction due to fewer possibilities for

593

recolonization and recovery (Culver et al., 2000; Pipan et al., 2010); they are

594

apparently more vulnerable to contaminants due to their narrow geographic ranges, 25

595

high patterns of endemism, and presumably low population size, low densities of

596

species, and low reproductive output (Di Lorenzo et al., 2014).

597

ii.

The low metabolic rates of subterranean organisms relative to phylogenetically

598

related surface species will likely influence toxicant uptake, metabolism and

599

depuration rates, which consequently may lead to different expression of toxicity

600

effects relative to surface species (Hose, 2005, 2007; Avramov et al., 2013; Di

601

Lorenzo et al., 2014, 2015b; Di Marzio et al., 2018). As a consequence, some authors

602

propose increasing the duration of acute toxicity tests with subterranean organisms

603

since a delayed manifestation of toxic effect is expected (Avramov et al., 2013; Hose

604

et al., 2016);

605

iii.

The biotic composition and structure, the abiotic conditions (and their interactions) in

606

subterranean ecosystems differ markedly from those in surface ecosystems.

607

Consequently, it is expected that subterranean biota will also have different

608

sensitivities towards stressors at the community level. The truncated trophic chains

609

and functional redundancy that are common in subterranean communities mean that

610

ecotoxicological testing using realistic approaches, specifically for subterranean

611

ecosystems are needed (Mösslacher and Notenboom, 1999; Daam et al., 2010).

612

There is presently no scientific evidence that subterranean organisms are overall more or less

613

sensitive than surface organisms to pollutants. This paucity of ecotoxicological data for

614

subterranean fauna has led to the use of sensitivity data for surface species as surrogate in the

615

development of environmental quality and protection thresholds. This approach is undesirable

616

because subterranean ecosystems have substantially different environmental conditions, food

617

chains and species traits to surface ecosystems. The development of standardized protocols

618

for testing and, with that, more data specific to subterranean taxa are needed to implement the 26

619

ERA process (Mösslacher and Notenboom, 1999; Hose, 2005; Humphreys, 2007; Daam et al.,

620

2010).

621

The priorities for supporting ERA in subterranean ecosystems should be to: 1) characterize

622

and quantify reliable data on contaminants’ concentrations found in the subterranean

623

ecosystems, including the emerging ones, which are currently out of regular monitoring

624

programs (Artigas et al., 2012); 2) prioritize substances for testing on subterranean species,

625

using existing information and sensitivity data on closely related surface species as surrogates

626

(Artigas et al., 2012, Di Lorenzo et al., 2019); and 3) select model species among the

627

subterranean ones, which should be a representative organism for each trophic level from both

628

compartment terrestrial and aquatic (EC, 2003). The criteria of species selection should be

629

based on knowledge of its morphology, biology. ecological function and routes of exposure

630

(Breithltz et al., 2006). Reporting information about the organism’s feeding behavior is also

631

desirable to enable comparisons between organisms from different trophic levels.

632

It is commonplace in ecotoxicological studies using subterranean species whose results are

633

compared to those for surface species exposed to the same toxicants. Such comparisons

634

provide useful context for the sensitivity of subterranean taxa particularly for comparisons

635

with model test species and those for which standardized protocols are available (e.g.

636

Notenboom et al., 1991; Avramov et al., 2013; Reboleira et al., 2013; Di Lorenzo et al., 2014;

637

Hose et al., 2016) or comparisons with field collected species from the same area of study or

638

other species more closely related (e.g. Bosnak and Morgan, 1981; Boutin et al., 1995; Di

639

Marzio et al., 2009; Di Lorenzo et al., 2014; Maazouzi et al., 2016; Raschmanova et al.,

640

2019). These comparisons have been inconsistent in demonstrating that either group is more

641

or less sensitive than the other. Bosnak and Morgan (1981), Boutin et al. (1995), Avramov et

642

al. (2013), Reboleira et al. (2013) and Maazouzi et al. (2016) observed that surface freshwater 27

643

species were more sensitive than groundwater species to the tested compounds; in contrast,

644

Barr (1976), Bosnak and Morgan (1981) and Di Lorenzo et al. (2014) observed the opposite

645

tendency. Hose et al. (2016) found little congruence between 96 h response data for Daphnia

646

magna and the response of the groundwater species. Interestingly, even subterranean

647

organisms from the same genus differed considerably in their sensitivity to the tested

648

compounds (Reboleira et al., 2013), and it was the species with the highest degree of

649

adaptation to subterranean life (P. lusitanicus) that was more sensitive to pollutants than the

650

species less adapted to subterranean ecosystems (P. assaforensis). For the temperature

651

tolerance, Raschmanova et al. (2019) observed that less adapted species to the subterranean

652

environment showed a wider range of temperature tolerance, but they did not observe

653

significant differences in UTL50 values between different ecological groups while the cold

654

resistance was significantly related to the body length within the same taxon (Collembola).

655

However, comparisons of surface and subterranean species remains confounded by

656

phylogenetic effects and paucity of data such that there are currently insufficient data for a

657

meaningful broad-scale comparison (Hose, 2005).

658 659

6. Future perspectives

660

The environmental and ecological characteristics of subterranean ecosystems demand

661

protocols to assess the impact of contaminants that are specific to the unique traits of the

662

biota. In parallel, a greater understanding of contaminant concentrations and fate in

663

subterranean ecosystems, and deeper knowledge of subterranean ecology are needed for a

664

holistic assessment of risk to subterranean ecosystems. Indeed, these represent the main

665

knowledge needs to progress a framework for environmental assessment of subterranean

666

ecosystems. The key steps to fulfill these knowledge needs are outlined below. 28

i)

667

Greater knowledge of subterranean ecology:

668

Key steps include:

669

•

Developing standardized methods for sample collection and monitoring to inform risk

670

assessments (Danielopol et al., 2003; Tomlinson et al., 2007; Humphreys, 2007, 2009;

671

Boulton, 2009; Larned, 2012);

672

•

Understanding the behavior and life cycles of subterranean fauna in order to develop laboratory cultures and enable chronic assays with sublethal end points (Larned, 2012);

673 674

•

Understanding the structure of subterranean trophic chains to underpin ecotoxicological

675

assays at the community level (e.g. understanding the role of subterranean fauna in the

676

cycling of carbon and nutrients) (Gibert and Deharveng, 2002; Korbel and Hose, 2011);

677

•

Increasing the knowledge of biodiversity, the relationship between biodiversity and

678

ecosystems functioning (Gibert et al., 2009), and the role of microorganisms and

679

invertebrates in the provision of ecosystem services (Larned, 2012; Hose and Stumpp,

680

2019).

681

ii)

Greater knowledge of contaminant fate and distribution in subterranean ecosystems

682 683

It is important to determine the current extent of contaminants in the subsurface, and to

684

predict the likely presence of contaminants into the future, which requires an understanding of

685

the fate and movement of contaminants in the subsurface (Stuart et al., 2012). Key steps

686

include:

687

•

Identifying actual and possible new subsurface pollutants, including the development

688

of new analytical techniques and technologies for in situ and real time analysis in

689

groundwater and sediments;

29

690

•

Increasing monitoring and assessment to determine the environmental occurrence of contaminants;

691 692

•

Characterizing the sources and identifying pathways of contaminant entry to the

693

subterranean environment, including connectivity with surface ecosystems (e.g.,

694

Bugnot et al., 2019);

695

•

Defining and quantifying the fate (persistence) and transport (dispersion) of contaminants in subterranean environments.

696 697

iii)

Identify potential ecological effects of contaminant exposure

698

In order to protect subterranean ecosystems and their fauna, it will be necessary to develop

699

new tools and/or adapt those that exist for other ecological domains to assess the chemical

700

and ecological state of subterranean ecosystems, considering also geological heterogeneity

701

and consequent regional endemicity patterns (Reboleira et al. 2013). Only when supported by

702

appropriate tools and data will robust environmental standards and legislation be established

703

to protect subterranean ecosystems. For example, the Water Framework Directive (EU, 2000)

704

and its Groundwater Directive (EU-GWD, 2006) are currently focused on setting the

705

threshold values for all pollutants that put groundwater bodies at risk of achieving a good

706

chemical state. However, to ensure the sustainability of subterranean ecosystems and

707

resources it is necessary to use an integrated approach that includes both the chemical and the

708

ecological status (Stuart et al., 2012). Key steps to achieve this include:

709

•

Setting appropriate ecotoxicological standards (protocols for culturing and protocols

710

for ecotoxicological assays) (Mösslacher and Notenboom, 1999; Gibert and

711

Deharveng, 2002; Danielopol et al., 2004; Hose, 2005; Humphreys, 2007; Di

712

Lorenzo et al., 2014, 2019), accounting for the specificities of these organisms and

30

713

following recommended guidelines as those for stygobiont crustaceans (Di Lorenzo

714

et al., 2019);

715

•

Establishing model or target subterranean species in order to generate a database of

716

chemical sensitivity to inform environmental risk assessments (Di Lorenzo et al.,

717

2019);

718

•

Working with realistic doses or mixtures and combining multilevel endpoints

719

(population, community and ecosystems) to assess the vulnerability of the

720

ecosystems from local to global scales;

721

•

Implementing an ERA for subterranean ecosystems that is based on relevant

722

scientific evidence and standard procedures for surface ecosystems (EMA, 2006,

723

2018), but modified in terms of requiring ecotoxicological data for at least three

724

different trophic levels to accommodate the simplified trophic chains in subterranean

725

ecosystems.

726

Finally, it is necessary to promote throughout society the importance and urgency of studying

727

and conserving subterranean resources to preserve their ecosystem services. It is also

728

necessary to create a science-policy dialogue to address management questions, ensure the

729

support and uptake of new research, and ultimately ensure the protection and sustainability of

730

subterranean ecosystems (Boulton, 2009).

731 732

Acknowledgements

733

This work was supported by a research grant (15471) from the VILLUM FONDEN.

734 735

Declarations of interest: none.

31

736 737

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1

Figure captions

2

Figure 1. Conceptual model of surface vs. subterranean ecosystems, over time and depth,

3

pointing out the simplified food web, generally lower organic matter content, higher relative

4

humidity and small temperature variation in the subterranean ecosystems.

5

Figure 2. Anthropogenic stressors in the subterranean ecosystems, with contamination sources

6

and threatened ecosystem services.

7

Figure 3. Maximum Concentration Detected (MCD) of contaminants in European

8

groundwater, data from the European Environmental Agency's Waterbase-quality (EEA,

9

2019). A) Metals; B) Pesticides; C) Fertilizers; D) Volatile Organic Compounds; and E)

10

Emerging contaminants. Asterisks points out the compounds that have been tested in

11

subterranean organisms.

12

Figure 4. World distribution of the subterranean species tested for stressors. Circles represent

13

species tested for contaminants sensitivity and squares represent species tested for upper-limit

14

survival temperature.

15

Figure 5. Graphical summary of the ecotoxicological endpoints for metal exposure on

16

groundwater species. End points shown: LC50 (Lethal Concentration, 50%) and IC25

17

(Inhibition Concentration, 25%). * no toxic responses for the maximum tested concentrations,

18

i.e., LC50 value is higher.

1

Highlights: •

Impacts of anthropogenic stressors on subterranean ecosystems are poorly known

•

Interdisciplinary research is needed for subterranean ecosystems’ sustainability

•

Subterranean biota is neglected in legal protection frameworks

•

Contaminants and temperature variation have pernicious effects on subterranean biota

•

Standardization of reliable testing protocols is urgently needed

Declaration of interests ☒ The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. ☐The authors declare the following financial interests/personal relationships which may be considered as potential competing interests: