Non-native species introductions and subsequent establishment continues to threaten global biodiversity (Seebens et al., 2017; Meyerson et al., 2019). The likelihood of a particular species successfully establishing and persisting in a novel environment is influenced by both abiotic and biotic factors (Byers, 2002; Pearson et al., 2018; Kinney et al., 2019). For example, evidence toward the environmental matching hypothesis suggests that when the climate of the incumbent ecosystem matches the climate of the native range, the invader has a higher potential to both persist and perform better than when the climate does not (Capinha et al., 2013; Iacarella et al., 2015). The community assemblage of the invaded system can also facilitate or impede establishment via biotic resistance (deRivera et al., 2005; Alofs and Jackson, 2014). Theoretically, areas with high species richness are more likely to have species that are strong competitors or, indeed, predators, which make the system less invadable than those with low species richness (Elton, 1958; Tilman, 1999). Although, in areas where species are functionally similar there is a likelihood that the analogous native species are outcompeted or niche excluded by the invader (Dick et al., 2017; Zeng et al., 2019).

Species interactions shape community dynamics through consumptive effects, but also in non-consumptive manners such as intraspecific and interspecific competition for resources such as food, shelter, and reproduction (Sih et al., 2010; Lopez et al., 2019; Mofu et al., 2019; Zeng et al., 2019). Biotic resistance can thus be exhibited in both consumptive and competitive manners, however, in freshwater systems biotic resistance is driven overwhelmingly by consumption (Alofs and Jackson, 2014). Aquatic environments exhibit a higher occurrence of generalist feeding and omnivory leading to a lack of intraspecific and interspecific competition (Alofs and Jackson, 2014). Moreover, dietary plasticity and frequency dependent predation (i.e., prey switching) are common traits of successful invaders as it allows persistence of species via trophic niche separation and capacity to consume new resources when one is over-exploited (Snyder and Evans, 2006; Olsson and Nyström, 2009; Havel et al., 2015). Often, successful invasive species also outcompete and competitively exclude native species for resources through agonistic interactions, though, aggressiveness does not always equate to resource holding potential (Camerlink et al., 2015; Lopez et al., 2019).

Freshwater crayfishes are highly successful invaders, having established widespread invasive populations. While the impacts, both positive and negative, have been reported for various crayfish species (see review by Lodge et al., 2012), the mechanisms that drive species persistence and increased impact are still somewhat unclear. The African continent is devoid of native crayfish species. This is potentially due to evolutionary competition with freshwater crabs (Ortmann, 1902; Lodge et al., 2012; Nunes et al., 2017a). Freshwater crabs of the genus Potamonautes are, however, present in almost all African freshwater habitats where they are trophically analogous to freshwater crayfish and provide essential nutrient cycling services (Hill and O'Keeffe, 1992; Dobson, 2004; Cumberlidge and Daniels, 2009; Peer et al., 2015). Potamonautid species typically exhibit high degrees of endemism and range restriction. These traits make them vulnerable to the impacts of habitat destruction and invasive species introductions (Cumberlidge and Daniels, 2009; Zeng and Yeo, 2018). Crabs and crayfish are polytrophic benthic omnivores that are both opportunistic scavengers and direct predators (Hill and O'Keeffe, 1992; Grey and Jackson, 2012). Due to the trophic similarities between crayfish and freshwater crabs it is likely that they will either provide an important component of biotic resistance or be competitively excluded by crayfish invasions (Lodge et al., 2012; Dick et al., 2017). Indeed, in Tanzania the red swamp crayfish, Procambarus clarkii (Girard, 1852) has replaced a native freshwater crab Potamonautes neumannii in many systems (Ogada, 2007), and has replaced Potamonautes loveni as the primary food source in African clawless otter diets in Lake Naivasha (Ogada et al., 2009). Similar trends are reported in Singapore where Australian redclaw crayfish Cherax quadricarinatus competitively excludes smaller native freshwater crabs from shelter resources (Zeng et al., 2019). In Cyprus the invasive P. clarkii shows shelter holding dominance over native freshwater crab Potamon potamios despite both species being equally aggressive (Savvides et al., 2015). Contrastingly, the European river crab Potamon fluviatile shows dominance in aggression and resource holding capacity toward the invasive P. clarkii but this may be driven by co-evolutionary history with the native Astropotamobius italicus crayfish (Cioni and Gherardi, 2004; Mazza et al., 2017). Again, there are similar reports that populations of native pseudothelphusid crabs persisting despite the C. quadricarinatus invasion in Mexico, likely due to the co-evolutionary history with native crayfish (Bortolini et al., 2007). Predicting the impacts of invasive crayfish species on native biota via consumptive and non-consumptive effects by way of comparing invasion histories can thus produce unequivocal conclusions due to geographical context.

Southern Africa is suffering from an over-invasion scenario by functionally similar crayfish species which have been introduced primarily through aquaculture ventures and the pet trade (Lodge et al., 2012; Russell et al., 2014; Nunes et al., 2017a,b; Weyl et al., 2020). Cherax quadricarinatus, the Australian redclaw crayfish and P. clarkii both have established invasive populations in South Africa (Nunes et al., 2017a,b). Cherax quadricarinatus is also present in Zambia (Nakayama et al., 2010; Nunes et al., 2016), Swaziland (Nunes et al., 2017a) and Zimbabwe (Marufu et al., 2018) while P. clarkii is invasive in Zambia, Uganda, Kenya, Egypt, Sudan, and Rwanda (Hobbs et al., 1989; Mikkola, 1996; Cumberlidge, 2009). Both species are likely to spread into ecologically and economically integral streams and wetlands, which will potentially threaten the stability of aquatic systems that provide refuge habitat for imperilled species such as Potamonautid crabs (Ahyong and Yeo, 2007; Belle et al., 2011; Nunes et al., 2017a,b; Zeng et al., 2019). To test whether there could be some degree of biotic resistance exhibited by African freshwater crabs, we measure and compare the maximum closing force of native Potamonautes perlatus to the invasive crayfish C. quadricarinatus and P. clarkii. Weapon performance can usually be assumed to be an honest signal and is correlated with aggression (Lappin and Husak, 2005; Wilson et al., 2007; Bywater and Wilson, 2012). We hypothesised that the invasive crayfish species would have a higher maximum closing force than P. perlatus. Closing force can indicate prey handling capacity (Meers, 2002; Miranda et al., 2016), but also, if closing force is related to morphology it can contribute to important signaling in agonistic contests (Wilson et al., 2007; Bywater et al., 2008; Bywater and Wilson, 2012). Thus, we also determined whether morphological and biological factors can be used to predict closing force.

There were significant differences in mass of each species (χ² = 39.03, df = 2, p < 0.001; Table 1, Figure 1), where P. clarkii weighed less than both C. quadricarinatus (z = 5.83, p < 0.001; Table 1, Figure 1) and P. perlatus (z = 4.11, p < 0.001; Table 1, Figure 1), however there was no difference between P. perlatus and C. quadricarinatus mass (z = 0.08, p = 0.93; Table 1, Figure 1).

Potamonautes perlatus had significantly stronger right chela closing force than left (Friedmans χ² = 15.05, df = 1, p < 0.0001; Supplementary Figures S1, S2). Contrastingly, both C. quadricarinatus and P. clarkii had significantly stronger left chela closing force (respectively: Friedmans χ² = 5.68, df = 1, p < 0.05, Friedmans χ² = 5.09, df = 1, p < 0.05; Supplementary Figures S1, S2). Female P. perlatus had a higher left and right chela closing force than males (Table 1, Figure 2) whereas, male C. quadricarinatus had a higher left and right chela closing force than females (Table 1, Figure 2). There were no sex differences in left or right chela closing force in P. clarkii (Table 1, Figure 2). Chelae length did not affect closing force in any species on either left or right chelae (left: χ² = 0.19, df = 1, p = 0.65; right: χ² = 0.83, df = 1, p = 0.36).

There was a significant interaction between species and sex on maximum closing force (Table 2, Figure 3; Supplementary Figure S3). Species and sex both had significant main effects on closing fsorce (Table 2, Figure 3; Supplementary Figure S3). Closing force increased significantly with mass for all species (Table 2, Figure 3; Supplementary Figure S3). Opposite trends in maximum closing force were seen between male and female P. perlatus and C. quadricarinatus, however there was no difference between male and female closing force in P. clarkii (p = 0.46). Female P. perlatus had a significantly stronger closing force than male P. perlatus, female C. quadricarinatus, and both sexes of P. clarkii (all p < 0.01). There was no difference between female P. perlatus closing strength and male C. quadricarinatus or between male P. perlatus and female C. quadricarinatus (both p > 0.05). However, male C. quadricarinatus had significantly higher maximum closing forces than male P. perlatus (p < 0.01). Male P. perlatus had significantly stronger maximum closing forces than both sexes of P. clarkii (all p < 0.01). There was no difference between maximum closing force of P. clarkii females and C. quadricarinatus females (p = 0.58).

Determining what makes systems resilient to biological invasion is a many faceted challenge (Holling, 1973; Havel et al., 2015). Nonetheless, by understanding basic differences in physical capacity it is possible to infer how performance can relate to species interaction outcomes within an invasion scenario (Griffen and Mosblack, 2011). We predicted that the invasive crayfish species would have a higher maximum closing force than native crabs, however, this proved to be unequivocal at least when considering animals of the same size. Given that there have been reports of invasive crayfish reducing abundance of trophically analogous decapods (Ogada, 2007; Ogada et al., 2009; Zeng et al., 2019), these results suggest that P. perlatus has the capacity to hold a competitive mechanical advantage over both invaders, but that this varies with sex. Unfortunately, our results are based on a small sample size of native P. perlatus which may confound results with regards to sex based differences. Nonetheless, there was a more even amount of males and females amongst smaller P. perlatus individuals (50–80 g) whereas the heavier individuals (>100 g) were predominantly female. Our results indicate that there are other mechanisms at play that may cause freshwater crabs to be competitively excluded by invasive crayfish species, rather than brute strength. Although, population size structure, growth rates, and maximum attainable mass of each species will affect competitive exclusion. For example, larger individuals will have a higher resource holding potential over smaller individuals of any species. Resultingly, it is possible to determine maximum chelae strength for these species through correlation with mass and sex but it is unclear as to whether chelae strength can actually be a predictor for resource holding potential or success in agonistic contests.

Closing force relates to ability of an individual to pinch down onto a subject. Whether this is for direct predation, during agonistic contests, reproductive purposes, or even to withstand abiotic disturbances such as high flow rates (Gherardi, 2002; Ion et al., 2019). It can be assumed that the larger the closing force the more damage may be conferred to the recipient despite high force conferring a high energy cost to the individual (Herrel et al., 1999; Wilson et al., 2007). Our results of chela closing force fall well within the range reported for P. perlatus (18–598 N) (Miranda et al., 2016), and for P. clarkii (males: 1.35 ± 0.41 N/g; females: 2.22 ± 0.89 N/g; see Supplementary Figure 3 for N/g results) (Claussen et al., 2007; Malavé et al., 2018). Similar to our results (Claussen et al., 2007) and Malavé et al. (2018) found sexual dimorphism in chela length where male P. clarkii had longer chela than females but this did not relate to closing force. This is likely related to reproductive activities and the cost of signaling during male-male contests (Stein, 1976). Contrastingly, P. perlatus females were stronger than males indicating a difference in resource holding potential between sexes and species. These differences may also be exacerbated in a natural setting when individuals are in or out of reproductive status. For example, female P. clarkii are more aggressive and more successful when they are maternal compared to non-maternal females and males (Peeke et al., 1995; Figler et al., 2005). Another native African crab species (Potamonautes sidneyi) had a significantly weaker closing force than P. perlatus (8–43 N) (Miranda et al., 2016). There is no prior published data on the closing force of C. quadricarinatus, but for comparative purposes the closing force of the largest terrestrial arthropod, the coconut crab (Birgus latro), is 29.4–1765.2 N (Oka et al., 2016), of which the lower ranges all overlap with the three species in the present study. In a first attempt to incorporate C. quadricarinatus into impact assessments, Zeng et al. (2019), show that larger bodied C. quadricarinatus have a competitive advantage over a native freshwater crab (Parathelphusa maculata) when competing for shelter space.

The presence of heterochely or “cutter vs. crusher” is well-established in marine decapods but it is less evident in freshwater crayfish species (Govind, 1989; Schenk and Wainwright, 2001; Lele and Pârvulescu, 2019). In essence it describes potential handedness between left and right chelae. Potamonautes perlatus had stronger right chela, whereas C. quadricarinatus and P. clarkii had stronger left chela. As claw length was not an effective predictor of strength it suggests a degree of ambidexterity between left and right chelae in all of these species as a response to likelihood of losing chelae during agonistic bouts (Kouba et al., 2011; Lele and Pârvulescu, 2019). The lack of strong morphometric predictors reinforces the concept of dishonest signaling in crayfish species (Wilson et al., 2007; Malavé et al., 2018). This combination of potential dishonest signaling and opposite trends in dominant chelae could be a factor in competition between crab and crayfish species despite P. perlatus females dominating mechanically. Potential for biotic resistance, either consumptive or competitive, is likely to be species specific and still further regulated by other biotic and abiotic parameters (deRivera et al., 2005). In this case, size mis-matches, differential spatial ecology, abundance, and type of resources present in a system are all possible factors that could be driving the likelihood of biotic resistance.

Freshwater crabs and crayfish are polytrophic keystone consumers that occupy an unusually large dietary niche breadth as a result of generalist feeding strategies (Jackson et al., 2014, 2016). Chelae are regularly used in decapod feeding to assist in subduing, capturing, holding, and manipulating resource items (Loya-Javellana et al., 1993; Mariappan et al., 2000). This is particularly important when considering durophagous feeding, as crushing predators are thus limited by their strength to process prey items such as snails, but also limited by the shell resistance and predator induced phenotype changes (DeWitt et al., 2000; Evers et al., 2011; Miranda et al., 2016). There is limited information on crush resistance of African gastropod species, but the invasive snail Tarebia granifera has a resistance of 100 ± 6 (mean ± se) N, while native Melanoides tuberculata has a resistance of 31 ± 4 N (Miranda et al., 2016). However, in its invasive range in Lake Malawi M. tuberculata has a crush resistance range of 18.63–94.73 N (Evers et al., 2011). Bulinus globosus and Bulinus nyassanus are native gastropods of Lake Malawi, which have crush resistance ranges of 2.29–4.79 N and 8.33–117.82 N, respectively (Evers et al., 2011). All three of the species represented in this study, besides female P. clarkii, have the capacity to handle all of these gastropod species. The relatively high crush resistance of the invaders suggests that this may facilitate their persistence in a system, however it should also be considered that if the southern African crayfish invasion persists there could be an invasion meltdown scenario where predation is concentrated on the native gastropods and facilitates population expansion of the invasive gastropods (Ricciardi, 2001; Simberloff, 2006). An invasion meltdown scenario may also be facilitated in the wild by Potamonautes sp. via differences in biotic resistance. Although, the cost-benefit of undertaking crushing activities for food, rather than reproductive efforts, should be investigated as it is likely that predators select for forage with low handling demands (Murdoch, 1969; Behrens Yamada and Boulding, 1998). The present study focuses on the relative differences in closing force with respect to resource utilisation and non-consumptive competition but direct predation by either crabs or crayfish upon heterospecific juveniles is also likely. Therefore, addressing actual contest outcomes and consumption rates between different sized individuals of each species would further our understanding of competitive interactions.

Handling vegetation and crushing prey items require differences in closing force and dentition patterns (Sibbing, 1991; Herrel et al., 1999). Thus, differences in closing force could also relate to niche separation between the species when they occur in sympatry, which could in turn facilitate species persistence of both natives and invaders. Procambarus clarkii exhibit this pattern of niche breadth reduction when found in sympatry with P. loveni, where they affect leaf litter breakdown due to direct consumption (Jackson et al., 2016; Nishijima et al., 2017), possibly related to differences in chela morphology (Sibbing, 1991). Further, P. clarkii is capable of exerting predatory pressure on planktonic prey items which do not need strong crushing capacity to handle (South et al., 2019). Little work has been completed on the diet and trophic niche of C. quadricarinatus in either its invasive or native range. Nonetheless, Marufu et al. (2018) found the main diet components of the Lake Kariba C. quadricarinatus population to be predominantly macrophytes, detritus and macroinvertebrates. In Lake Kariba the trophic niche of crayfish differed with size class, wherein macroinvertebrate consumption increased with size, which could potentially be due to the positive relationship between mass and closing force (Marufu et al., 2018). Comparative functional morphology of decapod chelae and feeding apparatus should thus be incorporated into invasion risk assessments as ecomorphology can help to predict impact (Nagelkerke et al., 2018).

Due to the complex nature of trophic interactions and food web structuring, particularly within a stochastic aquatic environment, it is difficult to determine mechanisms of biotic resistance in situ (Havel et al., 2015). Comparing species traits is a first step in assessing whether native species will exhibit some degree of either competitive or consumptive resistance toward invaders (Funk et al., 2008; Kumschick and Richardson, 2013; Zeng et al., 2015). The results presented here indicate that chelae closing force can be predicted by body mass and sex of the individual for all three decapod species but in order to correctly predict biotic resistance these must be validated further by assessing actual resource holding potential. Consequently, when trait based analysis should be complemented with other predictive assessments such as the comparative functional response and relative impact potential metrics (Dick et al., 2017; Dickey et al., 2018; South et al., 2019; but see Vonesh et al., 2017), but also with contest based experiments (Lopez et al., 2019; Zeng et al., 2019). Unfortunately, there is a severe paucity of data on the ecological impact of C. quadricarinatus but also on the basic ecology of Potamonautid crabs in southern Africa. Further, the specific dynamics of invasion and the recipient system can mediate trait expression in populations across the invasion gradient, whereupon crayfish at the invasion front can be more aggressive (Pintor et al., 2009), or they have smaller and less heavy chela as a response to reduced competition (Messager and Olden, 2019). Therefore, a considerable amount of baseline assessment (i.e., abundance, size structure, fecundity, distribution, and diet) is needed to be completed in order to effectively assess the risk that both invasive crayfish species pose toward functionally similar and ecologically important species such as freshwater crabs. Further, assessing actual interaction frequency and habitat or resource use overlap between different sized native and invasive decapod species would determine the potential degree of biotic resistance in the environment.

The raw data generated and used in the analysis are publicly available at https://doi.org/10.6084/m9.figshare.11993961.

This research was given ethics clearance by the Animal Ethics Subcommittee, Rhodes University (Ethics No. DIFS2718) and SAIAB Ethics Committee (#25/4/1/7/5_2018_06).

JS, TM, JM, and OW conceived the study. JM provided the equipment. OW provided the funding. JS, NT, and TM conducted the experiments. JS analysed the data and wrote the first draft. JS, TM, NT, JM, and OW all contributed toward manuscript editing and final manuscript.