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The unspoken truth about impacted rivers: Consequences and implications of barriersfor conservation of freshwater fish

Birnie-Gauvin, Kim

Publication date:2020

Document VersionPublisher's PDF, also known as Version of record

Link back to DTU Orbit

Citation (APA):Birnie-Gauvin, K. (2020). The unspoken truth about impacted rivers: Consequences and implications of barriersfor conservation of freshwater fish. Technical University of Denmark.

DTU Aqua National Institute of Aquatic Resources

Kim Birnie-Gauvin

The unspoken truth about impacted rivers Consequences and implications of barriers for conservation of freshwater fish

PhD Thesis

The unspoken truth about impacted rivers Consequences and implications of barriers for conservation of freshwater fish

PhD. Thesis by Kim Birnie-Gauvin National Institute of Aquatic Resources, Technical University of Denmark Section for Freshwater Fisheries and Ecology Submitted November 14, 2019

Supervisors: Professor; PhD Kim Aarestrup; Technical University of Denmark (main advisor) Senior research scientist; PhD Niels Jepsen; Technical University of Denmark (co-advisor) Professor; PhD Anders Koed; Technical university of Denmark (co-advisor)

Front cover image by Jack Perks ©

Either write something worth reading or

do something worth writing - Benjamin Franklin

(I tried to do both)

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Table of Contents

Acknowledgments .......................................................................................................................... 7

List of original manuscripts ........................................................................................................... 8

Summary .........................................................................................................................................10

Dansk résumé ............................................................................................................................... 12

Background ................................................................................................................................... 14

Freshwater Ecosystems ........................................................................................................... 15

Species distribution .............................................................................................................. 15

Migratory fishes ......................................................................................................................... 16

Salmonids ............................................................................................................................... 17

State of freshwater ecosystems .............................................................................................. 17

Threats to freshwater ecosystems: barriers ........................................................................... 18

Consequences of barriers .................................................................................................... 22

Re-establishing connectivity ................................................................................................ 28

Conclusions: management, implications and an outlook to the future ................................... 31

1. Fish passage is of little value if habitat changes are not addressed ................................ 31

2. Management needs to consider variation in migration timing .......................................... 32

3. Simultaneous restoration of passage and habitat yields massive reward ....................... 32

4. Multiple barriers have cumulative effects ............................................................................ 33

5. Fish are not meant to fit into equations ............................................................................... 34

Concluding remarks, and future directions ............................................................................ 34

Literature Cited ............................................................................................................................. 38

Manuscript I ................................................................................................................................... 51

Manuscript II .................................................................................................................................. 57

Manuscript III ................................................................................................................................. 63

Manuscript IV ................................................................................................................................ 69

Manuscript V ................................................................................................................................. 97

Manuscript VI .............................................................................................................................. 105

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PrefaceThis dissertation is submitted in partial fulfilment of the requirements for obtaining the

degree of Doctor of Philosophy (PhD) at the Technical University of Denmark (DTU). The

thesis represents work conducted during a three year scholarship in the period 2016-2019

at DTU, National Institute for Aquatic Resources (DTU Aqua), Section for Freshwater

Fisheries and Ecology. The project was funded by the Adaptive Management of Barriers in

European Rivers (AMBER – EU #689682), the Danish Rod and Net Fishing License Funds

and DTU.

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Acknowledgments I have a lot to be grateful for, and many people who deserve a special something here – if you’re uninterested, feel free to skip to the next section.

First and foremost, I wish to thank my primary supervisor Kim Aarestrup. No number of thank you’s will actually be enough to demonstrate how grateful I am. You’ve provided me with opportunities to learn and challenge myself that go above and beyond my expectations. Thank you. Thank you for your patience in dealing with my perfectionist being. Thank you for the debates and arguments, but also for the laughters.

To my co-supervisors, Niels Jepsen and Anders Koed, thank you for the chats, and the comic relief during meetings. Thank you for trusting me in carrying your expectations through.

A very very special thank you goes to Alice Manuzzi, Alan Le Moan and Romina Henriques. You guys have essentially kept me sane and alive during the past 3 years. You’ve given me a taste of the southern European culture by caring, partying, arguing, laughing and let’s not forget the food. You guys are the loudest people I’ve ever met, but you’re all amazing human beings, and I can’t imagine what life here would have been like without you. Thank you for the late evenings playing games, for the coffees, for the randomly thrown parties… Thank you.

A very big thank you goes to the technicians of our institute. Andreas Svarer, for putting up with me and my demands which usually pushed some boundaries, but also for dealing with all the boring work that was involved in monitoring barrier sites (though you were with me, so how bad could it really be?). A big thank you to Hans-Jørn A. Christensen and Michael Holm – our work together was not part of this thesis, but it was literally everything else. Thank you for coping with me, my clumsiness, my nosebleeds and my very spirited personality in the field. I hope all of you will continue to take me up on my wild ideas.

A sincere thank you to Henrik Baktoft and Hugo Flavio for all the stats help – lord knows I needed it. Martin L. Kristensen, thanks for all the collaborations, the jokes and entrusting me with some of your work – it’s been a blast.

À ma famille, merci du support inconditionnel durant tous mes aventures et mes idées folles.

It’s probably no secret that moving across the Atlantic Ocean to come spend a minimum of 3 years in a mostly unknown country, where people speak a weird ass language, is no easy task. Physically it’s easy – get on a plane. Emotionally though, nothing can really prepare you for it, but having been welcomed with open arms in this tiny middle-of-nowhere town that is Silkeborg has made all the difference. For this, I am grateful for this beautiful institute, and the beautiful souls that fill it with life and laughter every day. To all of you, thank you.

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List of original manuscripts

MANUSCRIPT I

Birnie-Gauvin, K., Aarestrup, K., Riis, T.M.O., Jepsen, N., Koed, A. (2017). Shining the light on the loss of rheophilic fish habitat in lowland rivers as a forgotten consequence of barriers and its implications for management. Aquatic Conservation: Marine and Freshwater Ecosystems, 27, 1345-1349. doi: 10.1002/aqc.2795. Status: Published.

MANUSCRIPT II

Aarestrup, K., Birnie-Gauvin, K., Larsen, M.H. (2018). Another paradigm lost? Autumn downstream migration of juvenile brown trout: evidence for a pre-smolt migration. Ecology of Freshwater Fish, 27, 513-516. doi: 10.1111/eff.12335. Status: Published.

MANUSCRIPT III

Birnie-Gauvin, K., Larsen, M.H., Nielsen, J., Aarestrup, K. (2017). 30 years of data reveal a dramatic increase in abundance of brown trout following the removal of a small hydrodam. Journal of Environmental Management, 204, 467-471. doi: 10.1016/j.jenvman.2017.09.022. Status: Published.

MANUSCRIPT IV

Birnie-Gauvin, K., Nielsen, J., Frandsen, S., Olesen, H.-M., & Aarestrup, K. Catchment-scale effects of river fragmentation: a case study on restoring connectivity. Status: in review in Journal of Environmental Management.

MANUSCRIPT V

Birnie-Gauvin, K., Candee, M. M., Baktoft, H., Larsen, M. H., Koed, A., & Aarestrup, K. (2018). River connectivity re-established: effects and implications of six weir removals on brown trout smolt migration. River Research and Applications, 34, 548-554. doi: 10.1002/rra.3271. Status: Published.

MANUSCRIPT VI

Birnie-Gauvin, K., Franklin, P., Wilkes, M., & Aarestrup, K. (in press). Moving beyond fitting fish into equations: Progressing the fish passage debate in the Anthropocene. Aquatic Conservation: Marine and Freshwater Ecosystems, 29, 1095-1105. doi: 10.1002/aqc.2946. Status: Published.

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List of additional manuscripts

Although these manuscripts were not included in this thesis, they reflect the results of additional work I have been involved in during my PhD.

Koed, A., Birnie-Gauvin, K., Sivebæk, F., & Aarestrup, K. (in press). From endangered to sustainable: multi-faceted management in rivers and coasts improves Atlantic salmon (Salmo salar) populations in Denmark. Fisheries Management and Ecology. doi: 10.1111/fme.12385

Birnie-Gauvin, K., Thorstad, E. B., & Aarestrup, K. (in press). Overlooked aspects of the Salmo salar and Salmo trutta lifecycles. Reviews in Fish Biology and Fisheries. doi: 10.1007/s11160-019-09575-x

Lennox, R.J., Paukert, C., Aarestrup, K., Auger-Methe, M., Baumgartner, L., Birnie-Gauvin, K., et al. (2019). 100 pressing questions on the future of global fish migration science, conservation, and policy. Frontiers in Ecology and Evolution. doi: 10.3389/fevo.2019.00286.

Kristensen, M.L., Birnie-Gauvin, K., Aarestrup K. (in press). Behaviour of veteran sea trout Salmo trutta L. in a dangerous fjord system. Marine Ecology Progress Series, 616, 141-153.

Birnie-Gauvin, K., Flavio, H., Kristensen, M. L., Walton Rabideau, S., Cooke, S. J., Willmore, W. G., Koed, A., & Aarestrup, K. (2019). Cortisol predicts migration timing and success in both Atlantic salmon and sea trout kelts. Scientific Reports, 9, 2422.

Birnie-Gauvin, K., Højrup, L. B., Kragh, T., Jacobsen, L., & Aarestrup, K. (in press). Getting cosy in freshwater: Assumed to be brackish pike are not so brackish after all. Ecology of Freshwater Fish. doi: 10.1111/eff.12460

Birnie‐Gauvin, K., & Aarestrup, K. (2018). A call for a paradigm shift: Assumed‐to‐be premature migrants actually yield good returns. Ecology of Freshwater Fish, 28, 62-68. doi: 10.1111/eff.12431

Kristensen, M. L., Birnie-Gauvin, K., & Aarestrup, K. (2018). Routes and survival of anadromous brown trout Salmo trutta L. post-smolts during early marine migration through a Danish fjord system. Estuarine, Coastal and Shelf Science, 209, 102-109. doi: 10.1016/j.ecss.2018.05.015

Birnie-Gauvin, K., Larsen, M. H., Thomassen, S. T., & Aarestrup, K. (2018). Testing three common stocking methods: differences in smolt size, migration rate and timing of two strains of stocked Atlantic salmon (Salmo salar). Aquaculture, 483, 163-168. doi: 10.1016/j.aquaculture.2017.10.021

Birnie-Gauvin, K., Tummers, J. S., Lucas, M. C., & Aarestrup, K. (2017). Adaptive management in the context of barriers in European freshwater ecosystems. Journal of Environmental Management, 204, 436-431. doi: 10.1016/j.jenvman.2017.09.023

Peiman, K. S., Birnie-Gauvin, K., Larsen, M. H., Colborne, S., Gilmour, K. M., Aarestrup, K., Willmore, W. G., & Cooke, S. J. (2017). Morphological, physiological and dietary covariation in migratory and resident adult brown trout (Salmo trutta). Zoology, 123, 79-90. doi: 10.1016/j.zool.2017.07.002

Birnie-Gauvin, K., Walton, S., Delle Palme, C. A., Manouchehri, B. A., Venne, S., Lennox, R. J., Chapman, J. M., Bennett, J. R., & Cooke, S. J. (2017). Conservation can inform threat assessment and recovery planning processes for threatened species. Endangered Species Research, 32, 507-513. doi: 10.3354/esr/00831

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Summary To serve the expanding land, water and energy needs of a growing human population, rivers have been altered extensively over the last centuries to build hydropower stations, weirs and reservoirs mainly for electricity, water security and irrigation, making freshwater the most threatened of ecosystems. Barriers come in many shapes and sizes, though most have had devastating effects on freshwater ecosystems by destroying habitats and critical spawning areas, hindering and delaying downstream and upstream passage for moving fish as well as leading to death and/or extinction for a range of species. While these effects have been investigated widely, there has been a focus on 1) large barriers, 2) the upstream passage of highly synchronized migratory species, 3) engineered solutions rather than removal, and 4) the effects at a local scale rather than catchment scale. As such, this thesis focuses on all of the ‘unspoken’ truths about barriers.

In the first manuscript (MS I), the loss of rheophilic habitat caused by barriers is highlighted. Stretches of river characterized by gravel-substrate, high-gradient, fast-flowing and highly-oxygenated water are inundated due to barriers, leading to slow-moving water and sedimentation, thus removing the most suitable habitat for salmonid spawning and early development. Using examples from catchments in Denmark, we demonstrate that multiple barriers have cumulative effects in terms of vertical and horizontal habitat loss, causing between 21 and 40% habitat loss. This habitat is crucial for many rheophilic species such as brown trout (Salmo trutta) and Atlantic salmon (Salmo salar). While both species are strong swimmers and are typically considered as highly synchronized migrants, recent evidence suggest that anadromous brown trout are highly variable in their migration timing. As such, migration timing was investigated in the second manuscript (MS II), where the assumption that all juvenile trout leave in the spring is questioned. Our findings suggest that trout also migrate in the autumn, and that these fish should be considered just as much as their spring counterparts with regards to barrier management. For example, practitioners should consider opening spillways during certain periods in the autumn, as is often done in the spring, to allow free passage. This is obviously not a perfect solution, but will certainly promote outmigration. A better solution to passage however is removal. In the third, fourth and fifth manuscripts (MS III, MS IV and MS V), the long-term effects of barrier removal are investigated. Manuscript III uses 30 years of data to investigate how removal has impacted the downstream and upstream spawning grounds and rearing areas, with focus on young of the year fish. Manuscript IV uses data collected over the last 25 years throughout an entire river system (including its tributaries) to study the effects of a creative pseudo dam removal. Manuscript V investigates the effects of six weir removals at a full river scale, with focus on the smolt run. All three studies demonstrate the benefits of removal, which are observed almost immediately. The habitats have returned to a more natural state, and the trout populations of river Gudenå (in the Vilholt region), river Kolding and river Villestrup are healthy. Lastly, the lessons learned from working with barriers, and especially the forgotten aspects and assumptions that render current practices ineffective, are discussed in the sixth manuscript (MS VI).

The results from this thesis are alarming when thinking of the overlooked effects of barriers (though perhaps not surprising), and yet positive when thinking of the effects that removing these barriers can have in reinstating lost habitats and fish communities. Management should seek to remove any barrier as a first option whenever possible, and use the results

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from this thesis as guidance to ensure that underlying assumptions on the way fish “should” behave and respond to barriers are questioned adequately.

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Dansk résumé For at kunne opfylde klodens stigende behov for fødevarer, vand og energi med en massivt voksende befolkning, er alle typer vandløb blevet ændret voldsomt i løbet af de sidste århundreder, ikke mindst gennem opstemninger med henblik på at lave elektricitet, sikre drikkevand og vand til kunstvanding. Disse opstemninger fungerer som barrierer for det dyreliv, der er i vandløbene. Spærringer findes i mange former og størrelser, men alle har negative virkninger på ferskvandsøkosystemer, blandt andet ved at ødelægge levesteder og kritiske gydeområder, hindre og/eller forsinke nedstrøms og opstrøms passage for fisk såvel som anden fauna og har endda ført til udryddelse af en række arter. På trods af mange undersøgelser af barrierer i vandløb gennem tiden, har der alt for ofte været fokus på 1) store opstemninger, 2) opstrøms passage af synkroniserede og vandrende arter, 3) modificeringer af opstemningerne frem for fjernelse og 4) virkningerne på lokal skala frem for hele vandløbssystemet. Som sådan fokuserer denne afhandling på alle de "uudtalte" sandheder om barrierer, med en særlig fokus på rheofile (strømelskende) arter (laksefisk) som model.

I det første manuskript (MS I) undersøges tabet af rheofilt habitat forårsaget af barrierer. De oprindelige strækninger opstrøms opstemninger, er normalt kendetegnet ved en høj vandløbsgradient og følgende hurtig vandstrøm, lav vanddybde og bund med grus substrat, samt gode iltforhold. Disse strækninger har ændret sig markant som følge af opstemningerne. Den hurtigstrømmende strækning erstattes af dybere og langsomt flydende vand, som igen leder til opbyggelse af fint sediment på bunden. Herved fjernes de mest velegnede habitater til gydning og opvækst for en række vigtige arter. Via eksempler fra vandområder i Danmark demonstrerer vi, at barrierer har kumulative virkninger og kan medføre mellem 21 og 40% habitat-tab, inden der overhovedet tages stilling til tab under vandring. Det tabte habitat er afgørende for mange reofile arter som ørreds (Salmo trutta) og Atlanterhavslaks’ (Salmo salar) trivsel og produktion. Begge arter er stærke svømmere og betragtes typisk som stærkt synkroniserede migranter, hvilket har ledt til enrække forvaltningstiltag skræddersyet til denne adfærd. Undersøgelser har dog indikeret, at blandt andet anadrome ørreder er meget mere variabel i migrationstiming end hidtil antaget. Derfor blev migrationstiming undersøgt i det andet manuskript (MS II), hvor antagelsen om, at alle juvenile havørred forlader åen om foråret, bliver belyst. Vores resultater viser, at rigtig mange ørreder vandrer nedstrøms til havet om efteråret, og at disse fisk bør tilgodeses lige så meget som forårs-artsfæller med hensyn til passage af opstemninger. Eksempelvis bør opstemnings-ejere åbne frislusen i visse perioder om efteråret (som det ofte gøres om foråret), for at tillade fri passage. Dette er selvfølgelig langtfra en ideel løsning, men vil helt sikkert tilgodese fiskenes udvandring. En bedre løsning på problemet er imidlertid en total fjernelse af opstemninger. I tredje, fjerde og femte manuskript (MS III, MS IV og MS V) undersøges de langsigtede virkninger af fjernelse af opstemninger. Her tales der altså ikke om at etablere en fisketrappe, et omløb, eller for den sags skyld et vandløb med unaturligt højt fald, men en fuldstændig fjernelse af opstemningerne og tilbageførsel af vandløbet til, stort set, dets oprindelige forløb og hældning. Manuskript III bruger 30 års data til at undersøge, hvordan fjernelsen af en opstemning har påvirket gydepladser og opdrætsområder for ørred. Fokus har specielt været på årets yngel (såkaldt YOY), både i selve området omkring opstemningen, men også på nedstrøms liggende gyde- og opvækstområder. Manuscript IV bruger data indsamlet i løbet af de sidste 25 år gennem et helt Å-system (inklusive dets tilløb) til at undersøge virkningerne af en kreativ ”pseudo

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fjernelse” af opstemning på antallet af både YOY og ældre ørred. Manuskript V undersøger virkningerne på et mere overordnet plan ved at undersøge effekten af fjernelse af seks opstemninger i det samme vandløb på smoltudtrækket. Alle tre studier viser fordelene ved en total-fjernelse af opstemningerne – fordele som desuden indtræder næsten umiddelbart efter spærringen er fjernet. Vandløbene er vendt tilbage til en mere naturlig tilstand, og tætheden af ørredyngel i Gudenå (ved Vilholt) og de øvre dele er Kolding Å, samt smoltudtrækket i Villestrup Å er mangedoblet. I det sjette og sidste manuskript (MS VI) diskuteres erfaringerne fra undersøgelserne af opstemninger, med særlig fokus på de glemte aspekter og antagelser, der risikerer at gøre den nuværende forvaltningspraksis ineffektiv.

Resultaterne fra denne afhandling er markante i forhold til de påviste negative virkninger af opstemninger. På samme tid er resultaterne positive fordi de viser de imponerende effekter som fjernelse af opstemninger har, i form af retablering af levesteder og fiskesamfund. I denne afhandling præsenteres der meget væsentlige grunde til at sigte mod en total fjernelse af opstemninger i vandløb. Denne afhandling kan bruges som en rettesnor for hvad man bør fokusere på i forvaltningen af vandløbenes fiskebestande og herigennem opnå en varig forbedring i vandløbsmiljøet generelt. Anbefalingen er klar; som den absolut første mulighed bør man forsøges at fjerne enhver opstemning i vandløbene. Der bør ligeledes være meget vægtige grunde hvis det vælges ikke at fjerne den. Påvirkningen af en enkelt opstemning rækker nemlig langt udover selve lokalområdet. Afhandlingen kan også bruges som en anvisning til hvilke effekter der kan forventes på bestandene af laksefisk ved en fjernelse af opstemninger.

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Background

Freshwater ecosystems and people have a complex relationship; freshwater fisheries are

an important component of the economy in many countries, but the sustainability of many

freshwater species is highly dependent on their ability to migrate to their spawning ground

as adults, and from their site of birth as juveniles. Though freshwater systems represent only

2.3% of the planet’s surface, 9.5% of all described species depend on them

(representing approximately 140,000 species; Balian et al. 2008; IUCN 2019). However,

these systems continue to be less studied, despite the fact that they are amongst the most

diverse, dynamic and complex ecosystems (Dynesius and Nilsson 1994). Freshwater

ecosystems are continuously challenged by severe threats which jeopardize their stability,

sustainability and biodiversity, making them the most imperilled ecosystems across the

world (Bruton 1995; Cowx 2002; Saunders et al. 2002; Reid et al. 2018).

Barriers – dams and weirs for example – pose a significant threat to freshwater ecosystems

and their inhabitants. While the effects of these artificial structures have been widely studied

on fish movement, less is known about the threats they may pose to other aspects of fish

lifecycles, the environment that surrounds them, and overall population dynamics. Though

it may be intuitive, we have little empirical evidence about the effects of removing these

barriers on fish.

As such, this dissertation addresses the following overarching questions:

• What are the overlooked consequences of barriers? (Manuscript I)

• What are the implications of barriers for fish with variable migrations? (Manuscript II)

• What are the effects of barrier removal at a local scale? (Manuscript III)

• What are the effects of barrier removal at a whole-system scale? (Manuscript IV and

V)

• How can we manage barriers more efficiently to benefit both fish and people?

(Manuscript VI)

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Freshwater Ecosystems

There exists great variability amongst freshwater ecosystems around the world, with which

both fish and managers must cope. In comparison with other regions of the world however,

European river systems consist of a wide range of shorter, slow-moving waters (Gough et

al. 2012), which will be the focus of this dissertation.

The flow of rivers is fed primarily by run-off from precipitations – from snow-melt in cold

temperate and high altitude regions, or from rain in warm temperate and tropical regions –

meaning that flow regimes vary depending on seasonality (Welcomme 1985). In Denmark,

most rivers are spring-fed with a higher proportion of groundwater influx. The type of terrain,

along with its vegetation, will impact the proportion of rain water and snow-melt that appear

as run-off, thus differences at a local scale can alter flow patterns differently in streams

located nearby (Welcomme 1985). As a result, smaller (or low-order) streams can become

altered quickly with rapid changes in flow regimes. As streams join to form larger rivers, the

flow is smoothed. In turn, the flow within rivers is related to the form of the basin and other

characteristics, such as the presence of forests and lakes, which tend to reduce flow

variability by distributing water more evenly, or storing water in the case of vegetation

(Welcomme 1985). The most distinctive difference between lowland streams and most other

watercourse is its limited gradient. In streams of lowland areas, as is the case in Denmark,

water flow tends to be slower, especially in the meandering parts of the watercourse

(Welcomme 1985). Due to their limited gradient however, these watersheds are often

regarded as less important (or even forgotten), with severe implications for the species

inhabiting them.

Species distribution

Given the variability of physical properties (i.e., depth, velocity, substrate etc.) and habitat

complexity along the length of streams and rivers, different species are bound to different

parts of the watercourse, where characteristics best suit their biology (Jackson et al. 2001;

Angermeier et al. 2002). The best suited habitat may change temporally and spatially

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however, depending on whether a fish is seeking food, spawning grounds or refuge habitat

(Northcote 1984), thus fish will often move between various types of habitats.

Migratory fishes

Migratory movements are amongst the most fascinating behaviours described to this day,

with examples spanning the entire animal kingdom. In fishes, migrations come in many

forms, such as the long-distance migrations of Atlantic salmon (Salmo salar) and European

eels (Anguilla Anguilla) between freshwater and marine environments (Klemetsen et al.

2003; Aarestrup et al. 2009), the long-distance migrations of bluefin tuna (Thunnus thynnus)

in marine environments (Block et al. 2005), the partial migration of brown trout (S. trutta,

Jonsson and Jonsson 2009) or the short-distance movement of cyprinids in and out of lakes

(Skov et al. 2008). These migrations are dependent on temporal, genetic, physiological and

environmental factors (McCormick et al. 1998; Lucas and Baras 2001; Aarestrup et al. 2002;

Lucas et al. 2009; Birnie-Gauvin et al. 2017a), though the underpinning mechanisms remain

largely unknown (Birnie-Gauvin et al. 2017b).

Many forms of migration in fishes have evolved over time. Two main types of migratory

movements involve freshwater: (1) potamodromy, which refers to movements occurring

solely in freshwater (Northcote and Hartman 2004), and (2) diadromy, which refers to a

transition between freshwater and salt water, and vice versa (McDowall 1995; Lucas and

Baras 2001). However, these movements are not as predictable and obligatory as first

thought (Lucas and Baras 2001; Aarestrup et al. 2018). For example, the proportion of

anadromous individuals in a brown trout (S. trutta) population is highly plastic; many

individuals will in fact assume residency in freshwater streams and never migrate to sea, a

phenomenon known as partial migration (Northcote 1992; Jonsson and Jonsson 1993;

Chapman et al. 2011). Other individuals are known to mature as male parrs or migrate into

lakes to feed rather than to sea.

Diadromy comes in two forms: anadromy and catadromy. Anadromous fishes are born

in freshwater and subsequently migrate to sea, where they undergo the majority of their

somatic growth, and then migrate back into freshwater to spawn are known as

anadromous fishes (McDowall 1997). By contrast, catadromous fish perform the majority

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of their somatic growth in freshwater, and spawn in marine waters (McDowall 1997).

These migratory movements are thought to have evolved either because fish exploit food

sources that are both temporally and spatially variable, they adapt to or avoid

environmental changes and predation, or they maximize their reproductive success

(Gross 1987; Ueda 2012). By doing so, fish also regulate population density and distribution

range (Ueda 2012). No matter the reason, the evolutionary basis of large-scale movements

is to ultimately gain fitness advantages by residing in a particular location, at a particular

stage (Northcote 1997; Dingle 2014). However, migrations are accompanied by several

costs and fish are constrained by limited energetic resources which must be allocated to

various fitness traits such as growth, metabolism, swimming, and avoiding predation

(Standen et al. 2002).

Salmonids

Salmonids are anadromous fishes, with a long history of being studied by researchers. Their

lifecycle has been studied and described extensively over the last century, though our

understanding of the mechanisms at play during the different stages remains surprisingly

limited. Adults migrate into freshwater to spawn. Eggs are buried in course gravel substrate.

Young alevins emerge from the eggs in the spring, and move up into the water column after

a short period to become fry and start feeding (Klemetsen et al. 2003). At the end of the first,

second or up to eighth summer, the smoltification process begins in order to prepare for

their future entry into saltwater (Björnsson et al. 2011). In the following spring, smolts will

migrate downstream, and out to sea where they will undergo most of their somatic growth.

They will later return to freshwater to spawn. However, over the last few decades, there

has been growing observations of the variation that exists within that lifecycle,

challenging the ‘textbook’ description of the salmonid paradigm. Given their

rather strict habitat requirements, salmonids are good indicators of system health.

Thus, we can use salmonids as indicators of habitat quality within a river. By measuring

salmonid numbers (either density, smolt run or spawning population), we can estimate the

type of habitat and habitat quality within the river, without measuring the habitat itself.

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State of freshwater ecosystems

Freshwater ecosystems provide services such as provisioning of clean water, recreation

and ecotourism for humans (also known as ecosystem services; Brauman et al. 2007;

Dodds et al. 2013), most of which are provided either directly or indirectly by the biota

(Loreau 2010). Freshwater also supports the highest concentration of biodiversity compared

to all ecosystems (Dudgeon et al. 2006). Consequently, any decline in freshwater species

richness and abundance is likely to have negative effects on the delivery of ecosystem

services (Ricciardi and Rasmussen 1999; Dudgeon et al. 2006; Macadam and Stockan

2015).

Figure 1. Living planet index decline of freshwater ecosystems between 1970 and 2012. The index has declined by 80% during that period, based on trends from 21271 populations of 4271 species.

Figure adapted from Reid et al. 2018.

In Europe, 531 freshwater fish species have been described and assessed, with more than

one third (37%) considered threatened (Freyhof and Brooks 2011). To put this number in

perspective, the European Red List has evaluated that 13% of birds and 15% of mammals

and dragonflies were threatened, but 44% of freshwater molluscs were, suggesting that the

current state of freshwater ecosystems is poor at best. In addition, an estimated 17% of

Europe’s freshwater fish populations are declining, and 76% remain unknown. This only

leaves approximately 7% of populations as stable.

18

Threats to freshwater ecosystems: barriers

Since the Bronze Age, deforestation and agricultural activities have heavily modified inland

waters by altering river discharge and sediment transport (Tockner et al. 2009). Overfishing

and species displacement have also been documented as early as the Roman times. While

these anthropogenic activities remain threats today, there is currently much focus on the

effects of climate change on the planet and its resources. While this is warranted, given that

climate change has had great impacts on the planet, freshwater ecosystems are already

greatly impacted by changes in human populations, and thus changes in water demands.

Modelling studies including outputs from climate models, water budgets and socioeconomic

data actually tell us that increasing water demands outweigh the effects of climate change

on global freshwater systems (Vörösmarty et al. 2000).

Over the course of history, humans have altered river systems in a variety of ways such as

for maintaining a trading route with other civilizations, or regulating water flow to generate

hydroelectricity or irrigation. Freshwater systems are threatened by a wide range of factors,

many of which can be traced back to humans.

19

Figure 2. Modelled predictions of the effects of climate change (top), population change (middle) and climate and population changes combined (bottom) on global freshwater supply. Blue

indicates wetter, green indicates no change, and red indicates drier. Figure from Vörösmaty et al. 2000.

20

Table 1. Threats to freshwater rivers and streams (modified from Suski and Cooke 2007). Highlighted in grey are threats especially relevant and discussed in this dissertation. Threat Severity Possible consequences

Introduction of alien species High Alien species outcompete endemics, endemic population declines/extinctions, increased spread of diseases and parasites

Barriers High Hindrance of passage for aquatic fauna, sedimentation, population isolation, loss of habitat, altered flow

Global warming High Alterations (increase) of water temperature, movement of species beyond natural range

Shoreline development High Habitat alteration, increased water temperature

Commercial fishing High Population decline, altered genetic flow

Recreational fishing Medium Population decline

Artisanal fishing Medium Population decline Flow regulation (reservoirs/abstraction) High Loss of habitat, altered hydrology

Fish farming Medium Loss of habitat, altered hydrology, eutrophication, introduction of diseases, parasites, species and genetic changes

Tourism High Pollution Land use (agriculture/forestry) High Loss of habitat and water quality,

eutrophication, acidification, sedimentation

Acidification High Release of metals, loss of habitat and water quality

Channelization High Loss of habitat, altered hydrology

Poor management High Piscicide use, exacerbation of all above threats

Whether for the purpose of acquiring food, generating energy, or for transportation,

extensive modifications have been made to rivers (Nilsson et al. 2005), with the distribution

and abundance of anthropogenic structures increasing dramatically since the 1950s (Ward

et al. 1999).

21

Figure 3. Number of large dams categorized by their purpose. Data from the International Commission on Large Dams (data extracted in January 2019).

Consequences of barriers

Damming is one of the most widespread anthropogenic modifications (Syvitski and Kettner

2011), and poses two major challenges to riverscapes and their fish inhabitants: (1) isolation

and/or eradication of fish populations (Moilanen and Nieminen 2002; McLaughlin et al. 2006)

and (2) loss of habitat (Malmqvist and Rundle 2002; Poff and Hart 2002; Nilsson et al. 2005;

Venter et al. 2006). In fact, barriers to fish movement are considered a major threat to the

abundance, distribution and life cycle of many fish species (Lucas and Baras 2001; Piecuch

et al. 2007; Lucas et al. 2009). These physical obstacles have been shown to hinder the

sustainability of fish populations with important long-term consequences (Levin 1974;

Merriam 1984; Ward and Stanford 1995). As an example, sea trout (Salmo trutta) smolts

losses at weirs associated with fish farms have been reported up to 94% in river Vejle,

Denmark. At the same location, 100% of salmon (S. salar) smolts were lost (Aarestrup and

Koed 2003). The extent of barrier effects varies from one location to the other, as well as

from dam to dam; trout smolt loss varied between 15% to 64% at four weirs in rivers Salten

and Mattrup, Denmark, with salmon doing generally poorer than trout (Aarestrup and Koed

22

2003). More drastic examples include an estimated 90-99% decline in European eel

(Anguilla anguilla) over the last few decades (ICES, 2017), associated largely to a high

mortality of silver eels at hydropower turbines as they migrate to marine waters (Pedersen

et al. 2012), as well as loss of habitat due to barriers.

Today, few (if any) river systems remain unaltered; they either have high-head (> 5 m) or

low-head (< 5 m) barriers. More than 45 000 dams above 15 m have been identified globally,

impounding 15% of the total river runoff annually (Avakyan and Iakovleva 1998; Gornitz

2001). A recent study suggested that 63%, 56%, 80% and 97% of rivers >1,000 km, 500-

1,000 km, 100-500 km and 10-100 km, respectively, were free-flowing (Grill et al. 2019).

This estimate assumed that rivers with a so-called connectivity status index of 95% or more

were free-flowing, and used national barrier databases which we know are not always

accurate (AMBER Atlas, data unpublished). Thus, it is reasonable to suggest that the

proportion of free-flowing rivers is significantly less than that estimated by Grill et al. (2019).

Figure 4. Barriers in Europe. Note: countries with no data are countries with no available databases, not countries with no barriers (figure from a preliminary version of the AMBER Atlas;

reproduction without permission is not permitted).

23

Whether small or large, these artificial structures obstruct passage through physical,

chemical and behavioural blockage or avoidance (Jepsen et al. 1998; Aarestrup and Koed

2003; Garcia de Leaniz 2008). Barriers fragment river catchments, and affect species’ ability

to disperse and migrate. In fact, fragmentation is thought to be the main cause of freshwater

species declines and extinction (WCD 2000) through the hindrance of both upstream

(Caudill et al. 2007) and downstream (Arnekleiv et al. 2007) movements. These alterations

have widespread effects on species presence and abundance both above and below dams

(Gehrke et al. 2002; Katano et al. 2006).

Table 2. Top 20 countries with the most dams according to the International Commission on Large Dams (2018). Note: ‘large dams’ are here defined as a dam with a height of 15 m or greater from lowest foundation to crest, or a dam between 5 m and 15 m impounding more than 3 million m3.

Country Number of barriers China 23 841 United States of America 9 265 India 5 100 Japan 3 118 Brazil 1 364 Korea 1 338 Canada 1 169 South Africa 1 112 Spain 1 063 Albania 1 008 Turkey 974 France 709 United Kingdom 593 Mexico 570 Australia 567 Italy 541 Iran 520 Germany 371 Norway 335 Zimbabwe 254 Denmark (ranked #80) 11 (4*)

*Note: While the International Commission on Large Dams (2018) notes 11 ‘large dams’ inDenmark, local experts estimate that no more than 4 such dams exist in Denmark.

In-river structures, in addition to blockage, may cause delays in migration for upstream

spawning migrations (e.g., Caudill et al. 2007) and downstream kelt and smolt migrations

(e.g., Acou et al. 2008). Delays may result from fish attempting to find a fishway entrance in

a somewhat chaotic flow, from repeated attempts in entering a fishway but being incapable

of completing it, or from fish being incapable of surpassing a barrier at a location where no

24

fishways are present (Pon et al. 2006; McLaughlin et al. 2013). While the extent of the delay

varies depending on location and species, no studies have reported a consistent absence

of migration delay at dam passing, further emphasizing that barriers never have a positive

effect on fish movements (McLaughlin et al. 2013). These delays also render individuals

more susceptible to predation by forming large aggregations of fish both below a barrier and

in the ponded zone, making for an attractive meal (Jepsen et al. 2000). For example, dams

have caused predation hotspots by predatory fishes and birds, targeting downstream

migrating salmonid smolts as they attempt to find their way past a barrier in the Columbia

River (USA; Schreck et al. 2006) and River Gudenaa (Denmark; Jepsen et al. 1998),

neotropical fish species in Tocantins River (Brazil; Agostinho et al. 2012), and western

minnow in Margaret River (Australia; Morgan and Beatty 2004).

Aggregations also facilitate the transfer of disease infections. Fish may sustain injuries as

they attempt to overcome these structures. Injuries may be internal or external and diminish

future survival. Surpassing barriers also increases energy expenditure which could

otherwise be allocated to other aspects of migration. This has several implications: (1)

individuals may arrive to spawning grounds with little energy for reproduction (Hinch and

Batty 2000; Caudill et al. 2007; Schilt 2007); (2) spawning success may be reduced through

a mismatch between hatching time and food availability (Cushing 1975); and (3) increased

mortality (Larinier 2001; Baisez et al. 2011; Roscoe et al. 2011). For example, the American

shad (Alosa sapidissima) population in the Connecticut River declined severely due to

delays in downstream migration which are thought to have reduced the post-breeding

survival of adults, thereby selecting against iteroparity (Castro-Santos and Letcher 2010).

25

Figure 5. Various types of barriers found across the world. A: Trend Dambrug, Trend, Denmark (now removed). B: Barrage de Poutès, Allier, France. C: unknown tidal weir near Warrnambool, southern Australia. D: Barrage de Vouglans, Jura, France. E: Gelstrup Dambrug, Nibe, Denmark (now removed). F: Linnmäe hydropower station, Jägala, Estonia. G: unknown hydropower station

under construction in Sa Pa, northern Vietnam. H: Sæby Vandmølle, Sæby, Denmark (no longer in use). I: Pitlochry Hydroelectric Dam and fish ladder, Pitlochry, Scotland.

Also related to delays is what we call a ‘fallback’. This occurs when a fish reverses its course

either before or after overcoming a barrier, typically in relation to using a fishway upon

upstream migration. A fish may become highly disoriented upon exiting the fishway and

move in the opposite direction, lack the motivation to continue to swim upstream or lack the

energy to pursue their migration (McLaughlin et al. 2013). Fallbacks can have important

ecological consequences: they present an additional delay during migration (Frank et al.

2009); they require a second attempt to overcome barriers whilst energy is depleted

(Reischel and Bjornn 2003); and they increase the probability of fish sustaining injuries. At

larger dams, thermal layering of the reservoir can lead to the development of temperature

gradients in fish passage structures. Evidence suggests that the larger the gradient

(between the fishway entrance and exit), the more likely are the fish to move back down a

fish ladder (Caudill et al. 2013).

26

One of the main issues with barriers is the continuous focus on fish passage. While it is true

that hindrance of passage at barriers has severe consequences on fish populations, the

presence of dams has had other debilitating effects such as ruining ideal spawning and early

growth habitats (Branco et al. 2012). For example, water flow changes from lotic (running

water) to lentic (standing water) (Guenther and Spacie 2006). Dams also cause

sedimentation, taking a watershed from having a mixture of rocks, cobbles and gravel as a

substrate to sand and mud. These factors lead to habitat loss, resulting in a decrease in

water quality and productivity (Carol et al. 2006), which consequently changes fish

assemblages (Tiemann et al. 2004; Gillette et al. 2005; Poulet 2007). Fish sometimes find

themselves in ecological traps, where they are somehow attracted or encouraged to

overcome a barrier, but the habitat upstream is not suitable for spawning. This could explain

why migrating adult Atlantic salmon in River Varde, Denmark, used a fishway to overcome

a barrier, but ‘overshot’ the potential spawning habitat, and moved back below the barrier to

spawn where the habitat was more suitable (Jepsen et al. 2003). Similar observations have

been made for Chinook salmon (Oncorhynchus tshawytscha) in the Columbia River, United

States (Boggs et al. 2004) and for Atlantic salmon in River Simojoki, Finland (Jokikokko

2002).

For salmonids (which are the focus of this dissertation), the main purpose of migrating back

to freshwater is for spawning. Salmonids rely heavily on the presence of rheophilic habitat

to achieve that goal, as they offer suitable habitat for laying eggs, egg emergence and early

growth and development. Because many freshwater fish species depend on migratory

movements to thrive and survive, be it short- or long-distance, they are particularly

vulnerable to population decline and risk of extinction (Maitland 1995; Suski and Cooke

2007). While much emphasis has been given to members of the Salmonidae family – given

their economic value and popularity among anglers – less consideration has been given to

the conservation of other species which are economically less influential, but remain

ecologically essential (Northcote 1998). There is growing evidence supporting long distance

movements in species considered as sedentary, with riverine fishes moving significantly

more than previously thought (Schlosser and Angermeier 1995; Jungwirth et al. 1998). Even

if few individuals successfully move from one population to another – crossing inadequate

and risky habitats by doing so – hindering the passage of these few individuals can have

27

substantial genetic and demographic consequences (Fausch 1995; Rieman and Dunham

2000). The presence of barriers and associated fishways may cause selectivity for certain

species, phenotypes or genotypes (Mallen-Cooper and Stuart 2007; Bunt et al. 2012). For

example, there is evidence to suggest that large fish with larger white muscle fibers were

more likely to successfully overcome a dam than smaller fish with smaller fibers (Volpato et

al. 2009). This selectivity is expected because individual fish will differ in their ability to swim,

orient and navigate, but favour residency over migratory phenotypes. The difficulty of

overcoming barriers increases the costs of migration, thus favouring resident

phenotypes, which are typically individuals of smaller size and lower reproductive potential

(Morita and Takashima 1998). Thus, by increasing their fitness, these individuals are

creating a local resident population which is more prone to extinction (Gyllenberg et al.

2002).

Re-establishing connectivity

The most recognized and arguably the most direct effect of barriers is blockage of migrants.

Hence, a great deal of research has been devoted to finding ways to develop, improve or

restore passage for migratory (and other target) fish species. An attempt to re-establish

connectivity has been made through the implementation of fishways (Poff and Hart 2002;

Gough et al. 2012; Tummers et al. 2016), though barrier removal appears to be a growing

practice. Fishways are designed with the objective of being ‘transparent’ such that they

facilitate fish movement around barriers without adverse effects on fitness (Castro-Santos

and Haro 2010), and have been around for at least a few centuries (first built more than 300

years ago in Europe; Clay 1995). They can be categorized as technical structures (e.g. Denil

fishways), nature-like structures (e.g. bypass channels) or special purpose structures (e.g.

eel ladders).

While the rate at which fishways are constructed is increasing, the success of these

structures remains low (Roscoe and Hinch 2010; Bunt et al. 2016). There are several

reasons as to why fishway success has been underwhelming. For example, fish swimming

performance has historically been used as a blueprint from which to design

structures. However, behaviour – how animals function within their physiological

limitations – is difficult to study,

28

and is likely different under laboratory conditions. Aspects of behaviour, such as motivation,

are particularly difficult to quantify and understand, though some progress has been made

in this field (Castro-Santos and Perry 2012).

Removal of in-river obstacles has obvious advantages over fishways, which merely mitigate

the fish passage issue, but do not resolve the problem at the source (Brown et al. 2013).

Barrier removal on the other hand, reinstates natural processes (Garcia de Leaniz 2008). It

is true that nature-like bypass channels, first introduced in the 1980s and 1990s in Europe

(Larinier 1998; Schmutz et al. 1998; Aarestrup et al. 2003), resemble a natural river stretch

with appropriate substrate, flow and gradient (Eberstaller et al. 1998) but all the issues

related to impoundments remain, even with this approach. To deliver any long-term benefits

on fish populations, fishways (no matter the type) must grant timely access to both upstream

and downstream moving fish without delays (Calles and Greenberg 2005).

Figure 6. Pre-removal and post-removal measurements. Density (per m2, with 95% confidence intervals) of young-of-the-year (YOY) and older year classes (OLD) in the first 50m upstream of weir position in A) river Trend (Salmo trutta), B) river Årup (Salmo trutta), C) river Idom (Salmo

salar), D) river Egtved (Salmo trutta), E) river Binderup (Salmo trutta), and F) river Ansager (Salmo trutta). Data was collected during my PhD as part of the AMBER project.

29

Figure 7. Conceptual diagram depicting the extent of barrier impacts through the various components of ecosystems, starting with direct impacts on habitat quality, and followed by more

downscale effects on species distribution, species abundance and overall ecosystem biodiversity.

30

Conclusions: management, implications and an outlook to the future

An estimated 15,000 fish species (out of 33,100 known species) depend on freshwater to

survive. Of these, 2,248 are considered threatened (IUCN 2019), suggesting that extinction

is a true possibility for many freshwater fish species (Pringle et al. 2000; Baras and Lucas

2001; Duncan and Lockwood 2001). This project investigated the effects of one of the major

causes for freshwater species declines and extinctions: barriers. Present in most of the

world’s rivers, barriers have had devastating effects on ecosystems and populations, often

established in the most crucial river stretches for migratory species or biodiversity hotspots

(Winemiller et al. 2016). Maintaining biodiversity of freshwater fish populations despite the

growing anthropogenic demands is necessary to ensure the continuity of the ecosystem

services they provide (Lynch et al. 2016).

The research included within this dissertation has led to five major conclusions, each of

which are described below.

1. Fish passage is of little value if habitat changes are not addressed

Manuscript I shines a light on a major issue with barriers: changes in habitat incurred

upstream of an obstacle. This includes changes in substrate, depth, flow, vegetation and

more. Barriers are typically constructed in areas of high slope gradient, thereby ruining the

ideal habitat that was previously found there, where spawning and early development

occurred. This loss of suitable conditions is true for both the horizontal and the vertical

dimension, and is a problem especially in lowland rivers where gradient is limited.

While barrier management has generally focused on getting fish past a barrier, the

conclusions of Manuscript I showcase the ineffectiveness (or perhaps uselessness) of

addressing passage issues without addressing the habitat loss – why move a fish upstream

if it is unsuitable there to spawn? The ultimate goal of migration for salmonids is to spawn

and contribute to the population. The provision of critical habitats for reproduction and

nurseries is necessary (Pompeu et al. 2012). However, if fish cannot find suitable spawning

grounds, and spawn in an area that is inadequate for the early growth phases, these

31

individuals are unlikely to survive and contribute to the next generation. This has important

implications for population sustainability, with declines as the unavoidable outcome.

2. Management needs to consider variation in migration timing

Even well-studied species have incomplete descriptions of their migrations (Winter et al.

2016). Of the 181 known fish species in Canadian freshwaters, 55% are considered

migratory (Lucas and Baras 2001), though we have a clear description of the migrations of

less than one third of them. Major connectivity problems are still present for the majority of

species in most regions (Foulds and Lucas 2013; Pelicice et al. 2015). Addressing these

concerns requires a considerable understanding of how species with lesser dispersing

capacities may overcome barriers (Gibson et al. 2005; Pépino et al. 2012; Silva et al. 2018).

With different species having different requirements for migration (and movements more

generally), different fish passage structures would be required for each species; this seldom

occurs in real life for well-studied species, let alone lesser known species. Manuscript II

investigates the proportion of brown trout smolts which migrates in the autumn, rather than

during the classically-described spring period. The findings suggest that close to 20% of the

“spring smolt class” migrates in the autumn, a significant proportion of the population.

These individuals do not have the same alleviations of barrier effects as their spring

counterparts, such as the opening of spillways during the peak smolt migration period.

From a management point of view however, it would be important to provide some relief for

these individuals as they likely contribute to population dynamics in a way that is not yet

understood.

Manuscript II illustrates the need to consider aspects of fish lifecycle that go beyond the

classical descriptions. It is likely that fish migrate more than we first thought, perhaps even

throughout the year.

3. Simultaneous restoration of passage and habitat yields massive reward

The lack of adequate habitats, or poor connectivity between these habitats, will likely result

in bottlenecks and population decline (Wilcox and Murphy 1985; Law and Dickman 1998).

This is especially true for species that require different habitats to complete their lifecycle,

32

making them particularly susceptible to the presence of obstacles in rivers (Amoros and

Bornette 2002; Mumby 2006). In some cases, anthropogenic barriers can reduce the fitness

of organisms so severely that the end result is population extinction (Poulsen 1935; Pringle

et al. 2000; Zabel and Williams 2002).

Manuscripts III, IV and V highlight the benefits of restoring both passage and

habitat through the removal of in-river structures. Taking such an (removal) approach

has had positive effects exceeding what had been hoped for in terms of

management objectives for population numbers, especially for fry, suggesting that

baselines can (and should) be shifted to aim for higher fish densities. All

manuscripts provide great examples of what can be accomplished when issues are

targeted at the source, with positive ramifications for conservation and

management, but also for local anglers and other stakeholders.

4. Multiple barriers have cumulative effects

Full appreciation of the adaptive value of fish movement requires consideration of migration

at a broader spatiotemporal level. Historically, there has been too much emphasis on the

upstream passage of adults, and too little on the downstream movements, including eggs,

juveniles and adults (Calles and Greenberg 2009; Bolland et al. 2012). This is especially

crucial in watersheds where more than one barrier is present. For example, fish that were

born farther upstream would need to successfully surpass multiple obstacles as smolts to

make it out to sea, with obvious cumulative effects. In such a case, management must

consider the cumulative (perhaps synergistic) consequences of having multiple barriers. On

the other hand, opting for full-river restoration where all barriers are removed (such as in

Manuscript V) results in what is possibly some of the best outcomes for conservation.

Manuscript V demonstrates the value of full river restoration where multiple barriers were

removed, with the spring smolt class increasing approximately 5-fold in 10 years.

33

5. Fish are not meant to fit into equations

The acceptance that freshwater ecosystems are in dire need of intervention, has led in part

to the realization of the importance of connectivity for migratory species. In most instances,

the solution has been to engineer a fishway, typically designed using equations (based on

lab-tests of fish swimming capabilities). This approach does not take into account the

different swimming abilities of various species (often focusing on strong swimming species

such as salmonids), their motivation and behaviour, or the variation across individuals or

states. It is rather an attempt to engineer a solution that will fit the most ‘typical’ fish. This

attempt only prolongs the existing problems that barriers cause, and certainly do not account

for hydrological and geomorphological properties of the river.

Manuscript VI highlights the most prominent issues with current fish passage practices from

a conservation and management perspective, and shines a light on the many ‘overlooked’

aspects of the current approach to managing barriers.

Concluding remarks, and future directions

My research during the past three years has led me to understand that three major obstacles

stand in the way of protecting watersheds in the context of barriers. While there are many

more issues which hinder proper management, these three are ones which cause major

drawbacks but that could easily be corrected.

Lack of consideration for small barriers

Low-head barriers are often thought of as having less overwhelming local effects than large

dams, and are believed to be passed more easily by fish (Welch et al. 2008). In other words,

smaller barriers have a lesser impact on fish populations individually than larger barriers

(Ovidio et al. 2007). However, their abundance is at the very least 2- to 4-fold that of large

barriers (Lucas et al. 2009), suggesting that their cumulative impacts are likely to be

significant on a larger scale (Lucas and Baras 2001; Cooke et al. 2005). Focusing on the

removal of such barriers will inevitably result in massive gains for fish populations.

34

Need for removal as a go-to tool

In 1998, Cowx and Welcomme wrote “Generally the soundest solution ecologically is to

remove structures as this not only restores longitudinal connectivity but can also lead to the

more general restoration of the habitat”. Yet in 2019, removal is rarely seen as a tool for

addressing the issues at hand. Barrier removal should be the go-to solution in any case, and

where removal is not a possibility, then other options can be explored.

There is currently a misconception in some literature that fish passage is a proven

technology; it is not. The success of fishways has in fact been underwhelming at best. While

it has been argued that the mediocre success of fishways is due to a lack of knowledge on

fish migratory requirements, it must also be acknowledged that most fishways are put in

place simply to “tick a box”. Legislation typically requires that owners and developers of

hydrodams must provide a fish passage structure, without the demands of post-

implementation monitoring. Furthermore, we currently lack an adequate framework for

evaluating the effectiveness of a fishway – how many fish is enough to call a fishway

successful?

Dam removal is typically considered a costly exercise, but it is arguably the most cost-

effective tool for long-term benefits. Fishways are expensive, and merely prolong the

problem. The costs of doing nothing (e.g. jobs, food, recreation, etc.) far too often exceed

the costs of removing a barrier or constructing a fishway (Silva et al. 2018). Removal almost

always leads to the immediate return and/or increase in fish populations (e.g. Elwha Dam,

Washington, USA; Tonra et al. 2015), and I therefore recommend that it become the first

option explored in any scenario and that exceedingly robust arguments would be needed to

maintain a barrier.

Inadequate management

It is not uncommon for restoration outcomes to fail to meet management objectives

(McLaughlin et al. 2013). Practitioners are increasingly coming to the realization that current

management is inadequate. We often lack the historical reference points to which we should

compare our data. More specifically, we come up with management goals without knowing

what the historical numbers were, leading to restoration efforts falling short of natural

35

conditions (Catalano et al. 2007; Benayas et al. 2009; Burroughs et al. 2010). In Europe,

the Water Framework Directive (WFD) – despite good intentions – has largely failed to

protect freshwater fish species. The directive allows for watersheds to be qualified as

“heavily-modified”, thereby exempting them from the specific goals set out to improve

freshwater ecosystems. The requirements for this qualification are simply too ‘plastic’, with

many stakeholders trying to get their respective rivers qualified as heavily-modified to avoid

complying with the legislation. Even rivers classified as “natural” still have barriers, which

have been maintained through the last two rounds of ‘water plans’. Legislation such as the

WFD needs to ensure the achievement of their environmental goals by establishing more

stringent demands and prohibit time exemptions (or apply severe penalties to member

states that do apply for time exemptions; Figure 8). The failure of a previous project

should not justify inaction (Silva et al. 2018).

Figure 8. European member states applying for time exemptions for implementing the EU Water Framework Directive. Figure adapted from WWF European Policy Office (February 2019).

Management also needs to contemplate broader aspects of migration. For example, sub-

lethal costs of fish passage are seldom considered in management (Silva et al. 2018).

Cortisol may become elevated following passage, and may reduce migration and

36

reproductive success, resource depletion or chronic stress (Burnett et al. 2014, 2017) as will

the increased risk of predation often associated with barriers.

Conclusion

In conclusion, it must be said that barriers have had horrendous effects on freshwater

ecosystems, not the least towards migratory fish populations. To that effect, it must also be

said that fish move more than is generally described, and thus the effects of barriers go far

beyond impacts on salmonids and other highly migratory species. The recommendations

resulting from this thesis can be summed in a few sentences. Remove all small barriers, as

most do not serve a purpose anymore, and stop considering poor excuses that compromise

the environment. Assume fish move and/or migrate more than the books say. And finally,

the only true management goal should be to reinstate habitat and connectivity in all rivers;

the approach used to do so should reflect that goal rather than giving way to exceptions.

37

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Manuscript I Shining a light on the loss of rheophilic fish habitat in lowland rivers as a

forgotten consequence of barriers and its implications for management

Published in: Aquatic Conservation: Marine and Freshwater Ecosystems, 27, 1345-1349 doi: 10.1002/aqc.2795

51

Received: 3 January 2017 Revised: 12 March 2017 Accepted: 16 April 2017

DO

I: 10.1002/aqc.2795

S HOR T COMMUN I C A T I ON

Shining a light on the loss of rheophilic fish habitat in lowlandrivers as a forgotten consequence of barriers, and itsimplications for management

Kim Birnie‐Gauvin1 | Kim Aarestrup1 | Thorsten M.O. Riis2 | Niels Jepsen1 | Anders Koed1

1DTU AQUA, National Institute of Aquatic

Resources, Technical University of Denmark,

Silkeborg, Denmark

2Limfjord Council, Nørresundby, Denmark

Correspondence

K. Birnie‐Gauvin, DTU AQUA, National

Institute of Aquatic Resources, Section for

Freshwater Fisheries Ecology, Technical

University of Denmark, Vejlsøvej 39, 8600

Silkeborg, Denmark.

Email: kbir@aqua.dtu.dk

Funding information

Danish rod and net fish licence funds;

European Union AMBER project (Adaptive

Management of Barriers in European Rivers).

Aquatic Conserv: Mar Freshw Ecosyst. 2017;1–5.

Abstract1. The majority of rivers around Europe have been modified in one way or another, and no longer

have an original, continuous flow from source to outlet. The presence of weirs and dams has

altered habitats, thus affecting the wildlife that lives within them. This is especially true for

migrating rheophilic fish species, which, in addition to safe passage, depend on gradient and

fast‐flowing waters for reproductive success and early development.

2. Thus far, research has focused on investigating the impacts of weirs and dams on fish passage,

with less attention paid to the loss of habitat entrained by such infrastructure. The loss of

rheophilic habitat is particularly important in lowland streams, where gradient is limited, and

dams and weirs can be constructed with less effort.

3. Denmark is considered a typical lowland country, where the landscape around streams and riv-

ers has been modified by agriculture and other human activities for centuries, leaving manage-

ment practitioners wondering how much change is acceptable to maintain sustainable fish

populations and fisheries practices.

4. With examples from Denmark, this paper attempts to conceptualize the loss in habitat as a

result of barriers in lowland streams and rivers, and the repercussions that such alterations

may have on rheophilic fish populations. Furthermore, the need for management to address

habitat loss and its related consequences concurrently with the improvement of fish passage

is emphasized.

KEYWORDS

catchment management, fish, hydropower, impoundment, indicator species, river, river

management, stream

1 | INTRODUCTION

The presence of barriers (such as weirs, dams and culverts) in rivers has

grown immensely in recent centuries. These barriers are most often

put in place to serve human needs, such as to generate electricity

(Welcomme, 1995), although fish farming, irrigation and flood control

are also common (Jungwirth, 1998; Jungwirth, Muhar, & Schmutz,

2000). When barriers were first established, the potential detrimental

impacts to the surrounding environment were not considered (Hunt,

1988), but it quickly became apparent that they had severe conse-

quences for river ecosystems and the organisms that live within them

(Aarestrup & Koed, 2003; Alexandre & Almeida, 2010; Dynesius &

wileyonlinelibrary.com

Nilsson, 1994; Junge, Museth, Hindar, Kraabøl, & Asbjørn Vøllestad,

2014; Koed, Jepsen, Aarestrup, & Nielsen, 2002).

Many countries lack a complete inventory of water barriers and

those that do typically register large barriers only (e.g. the United

States National Inventory of Dams for dams higher than 10 m). In

Denmark, the Ministry of Environment and Food has recently gener-

ated an inventory of barriers to implement the EC Water Framework

Directive (Council of the European Communities, 2000). Although

quite comprehensive, even this inventory is unlikely to account for

all Danish barriers, given that smaller weirs and especially culverts

often remain unregistered. While freshwater managers have remedied

some of the negative consequences of barriers associated with fish

Copyright © 2017 John Wiley & Sons, Ltd./journal/aqc 1

2 BIRNIE‐GAUVIN ET AL.

passage (e.g. through fish ladders, fish passes, etc.), most of the hab-

itat changes caused by damming are still present and thus still

threaten stream and river ecosystem sustainability. The need to take

action is pressing given that river ecosystems are in the poorest con-

dition of all ecosystems across the globe (WWF, 2016). To date, most

attention has been given to the impacts of barriers on fish passage

(both upstream and downstream movements; Aarestrup & Koed,

2003), and finding ways to establish minimum flows to sustain fluvial

habitat (Rood et al., 2005). While this approach has merit for manage-

ment, it ignores some basic problems: it does not account for the loss

of habitat in the ‘ponded’ zone that results from damming, and it

typically ignores the small‐scale migrations and movements of less

well‐known species (Larinier, 2001). Moreover, current management

schemes tend to neglect effects on other aquatic organisms, such as

plants and invertebrates, which are also affected by the presence of

obstacles (Merritt & Wohl, 2005; Palmer, Arensburger, Botts,

Hakenkamp, & Reid, 1995).

This paper briefly describes the important consequences of bar-

riers for rheophilic fish species (i.e. species that live in fast‐moving,

oxygen‐rich water), with greater focus on the quantity of habitat lost

owing to a loss in gradient, and lowland rivers and streams given

that gradient is a limiting factor for rheophilic fish reproduction

and development in such watercourses. We attempt to conceptual-

ize the loss in habitat as a result of barriers, and present a ‘quick

and dirty’ method that could be applied to management scenarios

that aim to restore the river continuum and natural habitats for

rheophilic fish species.

2 | HABITAT CHANGES AS ACONSEQUENCE OF BARRIERS

Barriers result in fragmentation and decoupling of hydrological, geo-

morphological and ecological aspects of a river, thereby modifying hab-

itat and restricting movement between them (Lucas & Baras, 2000;

McCluney et al., 2014; Nilsson, Reidy, Dynesius, & Revenga, 2005;

Poff et al., 1997; Ward & Stanford, 1983, 1995). Specifically, the

upstream section becomes a ‘ponded zone’ and the length of this zone

depends on the height of the dam and the watercourse gradient (Petts,

1984; Poff et al., 1997; Stanford et al., 1996; Figure 1). In turn, this

completely changes the river habitat upstream of the barrier, by

53

increasing the homogeneity of substrates and vegetation (Nilsson &

Jansson, 1995; Poff, Olden, Merritt, & Pepin, 2007), increasing the

depth, reducing current speed, reducing oxygenation, causing sedi-

mentation and changing water temperatures (Petts, 1984; Poff & Hart,

2002). The downstream habitat also becomes altered, but this paper

focuses primarily on the upstream geomorphological changes induced

by barriers.

3 | LOWLAND STREAMS AND RIVERS: CASESTUDIES FROM DENMARK

In lowland streams, the areas with relatively steep gradients are

selected preferentially for constructing barriers because of their

greater relative potential for energy (Hoffman & Dunham, 2007).

Damming effects also vary depending on the size of the watercourse

and the location of the dam. Generally, a dam located closer to the

source of a river will have fewer repercussions than one located

further downstream (Figure 1), because the gradient of the river is

typically greater in the upper regions, and therefore a smaller

proportion of the watercourse is affected by damming. Furthermore,

upstream parts of a river tend to be narrower than downstream

sections, thus the impacts of a dam are considerably lower when a

barrier is upstream (Figure 1), although it may still have important

consequences for local species.

In Denmark, a country consisting solely of lowland landscapes,

rivers are typically small, and have shallower gradients than those in

more mountainous countries. While a river in Norway, for example,

can easily provide a drop of 500 m, even the larger Danish rivers

typically rise below 100 m above sea level. Steep gradients are

therefore a limited resource in Denmark. Nonetheless, much of the

wildlife in Danish rivers relies on these scarce habitats (especially

rheophilic fish), making them especially important to protect. Within

lowland rivers, the areas where the gradient is (relatively) steep offer

greater potential for harnessing water power, often leading to the

establishment of more than a single dam throughout the river course.

For example, the River Grejs (Vejle, Denmark) runs for approx.

15 km, and has a total drop of 55 m from source to outlet, with 11

dams established by 1986.

An altered flow regime caused by dams affects the wildlife

present, typically reducing biodiversity (Bunn & Arthington, 2002;

FIGURE 1 Effects of dams on rivers.Conceptualized diagram of the effects of damson rivers showing two identical weirs (i.e.same stemmed height) (A and B). The pondedzone differs depending on the gradient of theriver. As the gradient typically decreases, andthe river size increases, from source to outlet,a similar sized weir closer to the outlet willhave a larger ponded zone, both in length andsurface area. downward‐pointing arrows (↓)represent a decrease

BIRNIE‐GAUVIN ET AL. 3

Power, Dietrich, & Finlay, 1996) and population size of migratory

species (Hubbs & Pigg, 1976; Zhong & Power, 1996). This is especially

true for rheophilic species (Hoffman & Dunham, 2007). Hence, an

increase in water level (i.e. increased depth) and a decrease in water

velocity may be used as indicators of the loss in geomorphological

variability and thus a river's ability to maintain biodiversity, as well as

a rough measure of potential rheophilic habitat loss. This is important

because a relatively large proportion of species that inhabit freshwater

streams require relatively fast‐flowing and oxygen‐rich water with

varied substrate conditions in order to thrive; the most common threat

to freshwater species (i.e. fish, amphibians, reptiles, mammals and

birds) is habitat loss and degradation from human activities (Freyhof

& Brooks, 2011).

Given the extent of dam establishment in some lowland rivers,

much of what used to constitute adequate habitats for these species

is no longer available. For example, indicator species of habitat

quality in Danish rivers, such as Atlantic salmon (Salmo salar) and

brown trout (Salmo trutta), spawn and grow (during early life‐stages)

in stretches where habitat is typified as riffle areas with gravel or

cobble substrate, and with low gradients (Gibson, 1993; Gibson,

Bowlby, & Amiro, 2008). Dammed rivers reduce the availability of such

stretches, and have been shown to reduce overall salmonid popula-

tions (Welcomme, 1985).

Recognizing the consequences of barriers on freshwater

ecosystems has led to the pursuit of mitigation strategies. For exam-

ple, some municipal and government agencies have put in place new

infrastructure to address environmental concerns (e.g. periodic high

flows, fish ladders; Auer, 1996). A common approach is the installa-

tion of nature‐like fish passes. These bypasses can be useful in

allowing fish to move upstream and downstream of a barrier (Calles

& Greenberg, 2005) but do not remedy the underlying habitat alter-

ations caused by barriers (Dadswell, 1996), and have been found to

have limited success (Bunt, Castro‐Santos, & Haro, 2012). Recent

evidence suggests that dam removal provides an efficient manage-

ment tool for ecological restoration of freshwater ecosystems

(reviewed in Bednarek, 2001), and should be considered where pos-

sible. In fact, regardless of how much knowledge there is on individ-

ual species, complete dam removal restores habitat quality, quantity

and connectivity, thus restoring previously lost habitat (Pess,

McHenry, Beechie, & Davies, 2008), enabling rheophilic fish popula-

tions to re‐establish and also enabling fish to migrate (both at small

and large scales).

TABLE 1 Conceptualizing rheophilic habitat loss. Using three Denmark riveof the river from source to outlet (m) was used as a proxy for vertical habitalength (km) was used as a proxy for horizontal habitat loss (%). This ‘quickmanagers with a low cost and effective method to gain a rapid overview ofimplementation of more effective management strategies

River (no. ofdams)

Total drop from sourceto outlet (m)

Summed drop frombarriers (m)

Verticlo

Villestrup (6) 22 8.8

Omme (14) 75 17.7

Gudenaa (7) 69 24.9

*Information not available given that the weirs and dams are too old to estimat

4 | CONCEPTUALIZING HABITAT LOSS:APPLICATIONS FOR MANAGEMENT

Table 1 provides data for three Danish rivers varying in size from 3 m

to 40 m in width and from 20 km to 149 km in length. These data

comprise the total drop from spring to outlet, the summed drop

resulting from barriers, the total length of the river, and the summed

length of the ponded zone, and were used as a rough estimate of

vertical and horizontal habitat loss. This information was chosen as it

is typically easy to obtain and to apply to management strategies.

We acknowledge that the habitat loss may not be proportional to the

loss in gradient (as this approach suggests). In fact, the relationship

between habitat loss and gradient is likely to be more complex,

especially if barriers are present further upstream, but this approach

has merit to address rapidly some of the present management

concerns.

This approach shows that a large proportion of the potential

rheophilic habitat is lost in the ponded zones (Table 1). The River

Gudenaa, the longest river in Denmark, was historically one of the

most important Danish rivers with large populations of anadromous

salmonids. It has seven barriers in the main stem predominantly for

hydropower generation, yielding a total relative loss of the potential

spawning and juvenile development habitat of 36% (Table 1). This loss

increases to approximately 60% if the upper 10% of the watercourse is

excluded where the river is narrow, the gradient is significantly

steeper, and salmon production is historically non‐existent. The smaller

rivers Villestrup and Omme, on the other hand, have barriers

established for fish farming or for driving old water mills, but

nonetheless result in a similar loss in habitat. Furthermore, this

estimated habitat loss is probably underestimated at fish farm sites,

because the stretch of the river between a weir and the outlet of a fish

farm is often several hundreds of metres apart, with very little water

flow during a large part of the year. The habitat quality in these

stretches is limited as a consequence of the reduced water flow alone,

but may also represent an area of high predation (Jepsen, Aarestrup,

Økland, & Rasmussen, 1998; Poe, Hansel, Vigg, Palmer, & Prendergast,

1991; Ruggerone, 1986).

The three rivers discussed above run mainly through agricultural

land. However, rivers running through urban areas may be subjected

to even more severe habitat loss (Birnie‐Gauvin, Peiman, Gallagher,

de Bruijn, & Cooke, 2016). For example, the River Mølleaa is approxi-

mately 13 km long, and flows through northern Copenhagen into the

rs, the ratio of the total drop as a result of barriers (m) to the total dropt loss (%). The ratio of the summed ponded zones (km) to the total riverand dirty’ approach to estimate habitat loss from barriers providesthe current state of freshwater streams and rivers, and may enable the

al habitatss (%)

Total riverlength (km)

Summed pondedzones (km)

Horizontal habitatloss (%)

40 20.0 5.8 29

24 55.0 11.35 21

36 149.0 ‐* ‐*

e accurately the length of ponded zones.

4 BIRNIE‐GAUVIN ET AL.

Øresund strait. The river has nine dams, which together remove an

estimated 75% of the river gradient. There is virtually no natural gradi-

ent left, and thus no adequate habitat for rheophilic species.

55

5 | CONCLUSIONS

The productive potential of rheophilic species in lowland rivers is

greatly reduced by the presence of dams and weirs. Typical manage-

ment interventions aim to address issues concerning fish passage, but

often omit any consideration of the habitat that has already been lost

as a result of barriers for which empirical data are lacking (Abell,

2002). Given the relatively limited gradient available in Danish rivers

(and in lowland rivers across the world in general) and the potential

habitat loss associated with the latter, the overall effects of water

barriers on habitat should be included in assessments of water-

courses. These actions should be undertaken concurrently with the

improvement of fish passage and other typical management‐related

challenges. To improve the state of regulated lowland rivers may

mean that many of these river obstacles need to be removed in order

to reinstate the former gradient and habitat, and at the same time re‐

establish faunal passage.

The purpose of this paper was to shine a light on a problem that is

often ignored in traditional fish management to this day: rheophilic

habitat loss resulting from barriers. Too often, the focus of manage-

ment is on fish passage alone, ignoring other important effects of dam-

ming. This may be particularly true for lowland rivers. Owing to the

number of dams and weirs in rivers across the world, we acknowledge

that acquiring a complete understanding of habitat loss and fish pas-

sage is a daunting task. However, if the majority of rheophilic habitat

is lost, improving fish passage may be pointless. We suggest, therefore,

the use of a ‘quick and dirty’method (Table 1) to evaluate the potential

loss in habitat as a result of barriers. This approach may provide man-

agers with an improved overview of the state of rivers, and allow for

better management strategies to be implemented. Further studies

should be undertaken to evaluate the validity of the approach.

ACKNOWLEDGEMENTS

This research was funded by the Danish rod and net fish licence funds

and the European Union (AMBER project).

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How to cite this article: Birnie‐Gauvin K, Aarestrup K, Riis

TMO, Jepsen N, Koed A. Shining a light on the loss of

rheophilic fish habitat in lowland rivers as a forgotten conse-

quence of barriers, and its implications for management.

Aquatic Conserv: Mar Freshw Ecosyst. 2017;1–5. https://doi.

org/10.1002/aqc.2795

Manuscript II Another paradigm lost? Autumn downstream migration of juvenile brown

trout: evidence for a pre-smolt migration

Published in: Ecology of Freshwater Fish, 27, 513-516 doi: 10.1111/eff.12335

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Ecol Freshw Fish 2017; 1–4  |  1© 2017 John Wiley & Sons A/S. Published by John Wiley & Sons Ltd

Accepted: 7 December 2016

DOI: 10.1111/eff.12335

L E T T E R

Another paradigm lost? Autumn downstream migration of juvenile brown trout: Evidence for a presmolt migration

The salmonidae life cycle is generally described as beginning in inland freshwater streams or rivers, where young fish (or parr) will spend the first 1–3 years of their lives. Parr then undergo extensive physiological and morphological transformation—also known as smoltification—to prepare for life at sea (Aarestrup, Nielsen, & Koed, 2002; Jonsson & Jonsson, 1993; Klemetsen et al., 2003). Once in marine waters, sal-monids undergo the majority of their somatic growth and return in freshwater to spawn, typically in late summer and autumn, after hav-ing spent 6 months to 3 years at sea (Davidsen et al., 2013).

While seaward migration of juvenile fish outside the peak spring period has been well documented in Atlantic salmon (Salmo salar; Cunjak, Chadwick, & Shears, 1989; Jonsson & Jonsson, 2014), few studies have published data to support an autumn outmigration in brown trout (Salmo trutta), despite the many observations of its oc-currence (e.g., Winter, Tummers, Aarestrup, Baktoft, & Lucas, 2016). In Atlantic salmon, this movement is considered a presmolt migration, perhaps as juveniles are entering estuaries, although they may not leave for sea until the following spring (Cunjak et al., 1989). For brown trout, it has been suggested that this movement is merely in- stream displacement or potadromous migration (Cucherousset, Ombredane, Charles, Marchand, & Baglinière, 2005). The underlying causes of these movements are unclear, and may require us to rethink what we have considered the paradigm of salmonid biology (coined by Gowan, Young, Fausch, & Riley, 1994). In this study, we aimed to characterise the length distribution and the sex of autumn- migrating trout in an attempt to better understand the underlying causes of autumn migration.

The River Lilleaa flows through agricultural land for approximately 35 km and is considered the most important sea trout tributary to River Gudenaa (K. Aarestrup, personal communication). A weir at Løjstrup Mill fish farm is situated two km upstream from the confluence with River Gudenaa, where a trap was erected to capture fish as they move downstream (Figure 1). Descending fish were caught in a wolf trap (8- mm grid spacing), fully covering fish passage within the stream. The trap typically operated in full unless river flow exceeded 5 m3·s−1. With larger flow, however, spill gates were open to allow excess water to exit the trap; thus, a portion of the fish were lost on days with ele-vated flow. The trap was running from 1 October to 29 October 2004 and was maintained daily. A subsample of migrating fish (n = 222) were measured for total length to the nearest ±5 mm on 18 October. Of these fish, 84 trout were randomly chosen to be sexed (through dis-section and genitalia identification). Data on water flow were obtained from a fixed data logger, located 100 m downstream of the trap.

We used a goodness- of- fit test to determine whether the sex ratio of migrating trout differed from 50:50 (male:female). A GLM

(generalised linear model) was used to assess whether the average length differed between males and females (data passed normality test and homogeneity of variance). All statistical analyses were performed in R version 3.0.1 (R Development Core Team 2013).

A total of 3,679 juvenile trout were caught between 1 October and 29 October 2004 (Figure 2). The start of capture coincided with the first major flow event. Unfortunately, the two major flow events (18 October and 22 October) necessitated the opening of the spill gates in order to release excess water, which enabled descending fish to escape capture. For this reason, we did not perform statistical analysis to in-vestigate the effects of water flow on downstream migration, although the data suggest a strong association with water flow (Figure 2). The average length of the migrating trout was 14.3 ± 2.7 cm, ranging from 9.0 to 21.0 cm (n = 222, Figure 3). The sex ratio was 68% females to 32% males and differed significantly from an even distribution (goodness- of- fit test, χ2 = 10.71, df = 1, p = 0.001, Figure 2). The aver-age length did not differ between males and females (GLM, F = 0.003, df = 1, p = 0.956).

A considerable proportion of the River Lilleaa brown trout popula-tion undergoes autumn downstream migration. Our findings suggest that autumn migration is likely greater than the number of fish captured in this study (i.e., >3,700) due to the occurrence of overflow events and the short trapping period. A similar study carried out within the same system in the year 2000 captured approximately 20,000 smolts during the peak spring migration (Aarestrup et al., 2002), providing ev-idence to suggest that autumn migration is not trivial (approximately 20% of spring migration numbers), and is likely more significant than first thought. The mean length of smolts in 2000 was similar to the length of the autumn- migrating trout captured in the present study (approx. 14 cm, K. Aarestrup, personal communication). Downstream migration of juveniles has been directly observed in other rivers such as the Imsa River, Norway (Jonsson & Jonsson, 2002). A recent telem-etry study by Winter et al. (2016) showed that autumn downstream migration of juvenile brown trout was evident in the Villestrup and Deerness streams of Denmark and England respectively. The evidence for autumn migration in S. trutta is growing and may have significant consequences on both stream and sea population dynamics (Winter et al., 2016), with important implications for management.

Autumn migrants may vary considerably from their spring coun-terparts in terms of their biology. For example, autumn migrants likely have lower gill sodium–potassium ATPase activity, as shown in juvenile Atlantic salmon (S. salar; Riley, Ibbotson, & Beaumont, 2009). However, we found that both the size range and the sex ratio were similar to that previously reported for spring smolts (Jonsson, 1985).

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2  | LETTER

In many ways, autumn migration resembles spring migration in River Lilleaa, suggesting that the movement of trout observed in autumn may be a presmolt migration and not just in- stream movements. The observed migrants with lengths ranging between 9 and 12 cm most likely contain both 0+ and 1+ fish, although we cannot draw that conclusion for certain given that we have no scales to confirm their age. Such individuals may in fact have reached the average smolt size

had they migrated the following spring. Consequently, we cannot exclude the possibility that population density may urge individuals to migrate sooner due to competition for proper habitat and food resources. Furthermore, we cannot exclude the possibility that this phenomenon only occurs in rivers exiting in low salinity areas.

The environmental and physiological factors that contribute to these migratory tactics are still poorly understood. Our study does

F IGURE  1 Map of the Lilleaa tributary and River Gudenaa with trap location

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F IGURE  2 Daily records of migrating trout during October 2004. The number of juvenile brown trout (Salmo trutta) captured daily during trap operation (n = 3679) accompanied by water flow. Opening of the water gate to allow excess water flow is represented by the thick black line

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|  3LETTER

not provide information on the final destination of the migrating trout, perhaps the Randers Fjord and Kattegat Sea, or somewhere between the trap and the sea. However, the habitat in this stretch is considered unfavourable for trout with its depth (~2 m), width (~30 m) and pres-ence of larger predatory fish such as pike (Esox lucius), suggesting that at least some of the trout detected at the trap migrated to sea although we cannot draw that conclusion for certain.

To this day, our knowledge of the movement ecology of salmonids remains incomplete (Dolinsek, McLaughlin, Grant, O’Connor, & Pratt, 2014), and may impact the effectiveness of management strategies currently in place which often relies on downstream fish passage data. Conservation decision- making processes also rely on population mea-surements and survival estimates (Mace et al., 2008). Failure to con-sider autumn migration in migration estimates may result in significant population and juvenile to adult survival underestimates and thus have important conservation and management consequences (e.g., when regulating fish passage opening periods and spillway release at weirs and dams). The occurrence of autumn migration may also affect age determination using scales which relies on the width of growth rings to estimate time spent at sea and assumes that the first time fish enter marine waters is in the spring (Bagenal, 1974), which may affect esti-mates of population and age dynamics. We urge the development of more telemetry studies to evaluate the extent of this autumn migra-tion, as it may have important consequences on salmonid management in streams of many regions of Europe. We further emphasise the need to consider this autumn- migratory phenotype as part of the salmonid paradigm (Gowan et al., 1994), as it may play an adaptive role in salmo-nid life- history strategies due to changing environmental conditions.

ACKNOWLEDGEMENTS

This study was funded by the Danish rod and net fish licence funds.

Kim Aarestrup1

Kim Birnie-Gauvin1

Martin H. Larsen1,2

1DTU AQUA, National Institute of Aquatic Resources, Section for Freshwater Fisheries Ecology, Technical University of Denmark, Silkeborg,

Denmark2Danish Centre for Wild Salmon, Brusgårdsvej 15, DK-8960 Randers SØ,

Denmark

CorrespondenceKim Aarestrup, DTU AQUA, National Institute of Aquatic Resources,

Section for Freshwater Fisheries Ecology, Technical University of Denmark, Silkeborg, Denmark.

Email: kaa@aqua.dtu.dk

REFERENCES

Aarestrup, K., Nielsen, C., & Koed, A. (2002). Net ground speed of down-stream migrating radio- tagged Atlantic salmon (Salmo salar L.) and brown trout (Salmo trutta L.) smolts in relation to environmental fac-tors. Hydrobiologia, 483, 95–102.

Bagenal, T. B. 1974. The ageing of fish: The proceedings of an international symposium. Old Woking, Surrey, UK: Unwin Brothers Ltd.

Cucherousset, J., Ombredane, D., Charles, K., Marchand, F., & Baglinière, J.-L. (2005). A continuum of life history tactic in a brown trout (Salmo trutta) population. Canadian Journal of Fisheries and Aquatic Sciences, 62, 1600–1610.

F IGURE  3 Length (n = 222) and sex (n = 84) distribution of a subsample of juvenile brown trout (Salmo trutta) captured on 18 October 2004

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4  | LETTER

Cunjak, R. A., Chadwick, E. M. P., & Shears, M. (1989). Downstream move-ments and estuarine residence by Atlantic salmon parr (Salmo salar). Canadian Journal of Fisheries and Aquatic Sciences, 46, 1466–1471.

Davidsen, J. G., Rikardsen, A. H., Thorstad, E. B., Halttunen, E., Mitamura, H., Præbel, K., … Næsje, T. F. (2013). Homing behaviour of Atlantic salmon (Salmo salar) during final phase of marine migration and river entry. Canadian Journal of Fisheries and Aquatic Sciences, 70, 794–802.

Dolinsek, I. J., McLaughlin, R. L., Grant, J. W., O’Connor, L. M., & Pratt, T. C. (2014). Do natural history data predict the movement ecology of fishes in Lake Ontario streams? Canadian Journal of Fisheries and Aquatic Sciences, 71, 1171–1185.

Gowan, C., Young, M. K., Fausch, K. D., & Riley, S. C. (1994). Restricted movement in resident stream salmonids: A paradigm lost? Canadian Journal of Fisheries and Aquatic Sciences, 51, 2626–2637.

Jonsson, B. (1985). Life- history patterns of freshwater resident and sea- run migrant brown trout in Norway. Transactions of the American Fisheries Society, 114, 182–194.

Jonsson, B., & Jonsson, N. (1993). Partial migration: Niche shift versus sexual maturation in fishes. Reviews in Fish Biology and Fisheries, 3, 348–365.

Jonsson, N., & Jonsson, B. (2002). Migration of anadromous brown trout Salmo trutta in a Norwegian river. Freshwater Biology, 47, 1391–1401.

Jonsson, B., & Jonsson, N. (2014). Time and size at seaward migration influ-ence the sea survival of Atlantic salmon (Salmo salar L.). Journal of Fish Biology, 84, 1457–1473.

Klemetsen, A., Amundsen, P. A., Dempson, J. B., Jonsson, B., Jonsson, N., Oconnell, M. F., & Mortensen, E. (2003). Atlantic salmon Salmo salar L., brown trout Salmo trutta L. and Arctic charr Salvelinus alpinus L.: A review of aspects of their life histories. Ecology of Freshwater Fish, 12, 1–59.

Mace, G. M., Collar, N. J., Gaston, K. J., Hilton-Taylor, C., Akçakaya, H. R., Leader-Williams, N., … Stuart, S. N. (2008). Quantification of extinction risk: IUCN’s system for classifying threatened species. Conservation Biology, 22, 1424–1442.

R Core Team. (2013). R: a language and environment for statistical comput-ing. R Foundation for Statistical Computing, Vienna, Austria.

Riley, W. D., Ibbotson, A. T., & Beaumont, W. R. C. (2009). Adult returns from Atlantic salmon, Salmo salar, parr autumn migrants. Fisheries Management and Ecology, 16, 75–76.

Winter, E. R., Tummers, J. S., Aarestrup, K., Baktoft, H., & Lucas, M. C. (2016). Investigating the phenology of seaward migration of juvenile brown trout (Salmo trutta) in two European populations. Hydrobiologia, 775, 139–151.

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Manuscript III 30 years of data reveal a dramatic increase in abundance of brown trout

following the removal of a small hydrodam

Published in: Journal of Environmental Management, 204, 467-471 doi: 10.1016/j.jenvman.2017.09.022

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lable at ScienceDirect

Journal of Environmental Management 204 (2017) 467e471

Contents lists avai

Journal of Environmental Management

journal homepage: www.elsevier .com/locate/ jenvman

Research article

30 years of data reveal dramatic increase in abundance of brown troutfollowing the removal of a small hydrodam

Kim Birnie-Gauvin a, *, Martin H. Larsen a, b, Jan Nielsen a, Kim Aarestrup a

a DTU Aqua, Section for Freshwater Fisheries and Ecology, Vejlsøvej 39, 8600 Silkeborg, Denmarkb Danish Centre for Wild Salmon, Brusgårdsvej 15, 8960 Randers, Denmark

a r t i c l e i n f o

Article history:Received 22 June 2017Received in revised form29 August 2017Accepted 7 September 2017

Keywords:ConservationDamsFish passageMigrationPopulationSalmonidae

* Corresponding author.E-mail address: kbir@aqua.dtu.dk (K. Birnie-Gauvi

http://dx.doi.org/10.1016/j.jenvman.2017.09.0220301-4797/© 2017 Elsevier Ltd. All rights reserved.

a b s t r a c t

Humans and freshwater ecosystems have a long history of cohabitation. Today, nearly all major rivers ofthe world have an in-stream structure which changes water flow, substrate composition, vegetation, andfish assemblage composition. The realization of these effects and their subsequent impacts on populationsustainability and conservation has led to a collective effort aimed to find ways to mitigate these impacts.Barrier removal has recently received greater interest as a potential solution to restore river connectivity,and reestablish high quality habitats, suitable for feeding, refuge and spawning of fish. In the presentstudy, we present thirty years of data from electrofishing surveys obtained at two sites, both prior to andfollowing the removal of a small-scale hydropower dam in Central Jutland, Denmark. We demonstratethat the dam removal has led to a dramatic increase in trout density, especially in young of the year.Surprisingly, we found that this increase was not just upstream of the barrier, where the ponded zonepreviously was, but also downstream of the barrier, despite little changes in habitat in that area. Thesefindings suggest that barrier removal may be the soundest conservation option to reinstate fish popu-lation productivity.

© 2017 Elsevier Ltd. All rights reserved.

1. Introduction

Obstacles within watercourses, such as dams and weirs, havebecome pervasive in today's freshwater ecosystems. Beginning inthe tenth century, humans have modified rivers to operate mills,net fish as a food source, navigate to trade with foreign countries,generate energy and regulate water (Baxter, 1977; Dudgeon, 1992;Northcote, 1998; Downward and Skinner, 2005; Nützmann et al.,2011). Today, scarcely any river systems remain unaltered byanthropogenic structures (Hall et al., 1991; Morita and Yamamoto,2001).

The impacts that dams have had on freshwater ecosystems areconsiderable; alterations to the physical and chemical characteris-tics of the water and surrounding landscapes has resulted in theincrease of homogeneity and a decrease in suitable habitat formany species, including the loss of low-water spawning andnursery habitats for salmonids and lampreys in ponded zones(Baxter, 1977; Jungwirth et al., 2000; Birnie-Gauvin et al. in press);and interference with one or more stage in the life cycle of many

n).

64

fish species has led to changes in fish assemblages (Lucas and Baras,2001). For example, brown trout (Salmo trutta) and Atlantic salmon(S. salar) smolts showed significant delays and increased mortalitywhen released upstream of a small weir in comparison to in-dividuals released downstream of the barrier (Aarestrup and Koed,2003). Furthermore, the natural flow patterns of regulated rivers,which provide important cues for fish migrations, have beenaltered extensively, thereby also reducing biodiversity (Bunn andArthington, 2002). This reduction in biodiversity and populationnumbers is further exacerbated by an increased mortality ofmigratory fish in reservoirs (or ponded zones) formed by dams(Jepsen et al., 1998). Fish will often accumulate in these pondedzones, as well as just downstream of a dam (Koed et al., 2002),making them more susceptible to predation by other fish and toexploitation by fisheries (Poe et al., 1991; Lucas and Baras, 2001).Taken together, the construction of dams and weirs is estimated toaccount for 55e60% of the known causes leading to freshwater fishendangerment (Northcote, 1998). To aggravate their status, fresh-water species are already considered more imperiled than terres-trial species (McAllister et al., 1997; Ricciardi and Rasmussen,1999),requiring us to take action.

The recognition of the negative impacts of barriers in the last

Abbreviations

YOY young of the yearOLD older fish

K. Birnie-Gauvin et al. / Journal of Environmental Management 204 (2017) 467e471468

few decades has led to the quest for solutions that would enablesafe passage. For example, many hydrodams in the United Stateshave adopted the policy of manually trapping and moving fishpassed dams (Cada 1998). In other cases, fish passes, such as fishladders, fish elevators or nature-like fish passes, have been imple-mented (see Jungwirth et al., 1998 for review). Despite these effortshowever, the efficiency of fish passage facilities remains under-whelming in many cases. In the River Gudenaa, Denmark, theTangeværket Dam has resulted in the extinction of Atlantic salmonand the near-elimination of upstreammigrating sea trout (S. trutta),despite the presence of a fish ladder (Aarestrup and Jepsen, 1998).

Though larger obstacles are viewed as having more significantconsequences, smaller barriers such as weirs are more common(estimated two-to four-fold; Lucas et al., 2009). On a large scale,their cumulative effects are likely to be significant (Cooke et al.,2005), though these low-head barriers continue to be less studied(Lucas and Baras, 2001). Nomatter the barrier size however, barrierremoval is presumably the most appropriate solution (Cowx andWelcomme, 1998). It (1) restores longitudinal connectivity, (2) re-stores the natural habitat (including physical and chemical prop-erties), and (3) enables safe fish passage. Despite the recognitionthat removal is likely the soundest of all conservation options since1998 (by Cowx and Welcomme), relatively few studies haveexamined the consequences of barrier removal (but see Bednarek,2001 for review on ecological effects), especially in the context ofsmaller obstacles and over long timescales. The recovery responseof fish populations and communities to removal remains largelyundocumented (Doyle et al., 2005) making it difficult to makepredictions and influence decisions made at the management level.Existing recommendations include viewing small barrier removalsas opportunities to educate ourselves on the impacts beforecontemplating large barrier removals, which are likely accompa-nied by greater consequences (Doyle et al., 2003). The few studiesthat have examined the effects of barrier removal on fish assem-blage and distributions have been carried over relatively shortperiods of time, but all indicated or predicted positive impacts ofremoval on native species (e.g., Catalano et al., 2007; Pess et al.,2008; Burroughs et al., 2010; Hitt et al., 2012). Here, we present30 years of data on brown trout numbers both before and after theremoval of a small hydropower dam (Vilholt, Central Jutland,Denmark). Such temporal data on the subject has never beenavailable prior to this study (that we know of), making it the first ofits kind.

2. Methods

2.1. Study site

River Gudenaa is one of the largest rivers in Jutland, Denmark,running for approximately 149 km before entering the RandersFjord. In 1866, the Vilholt hydropower dam (Vilholt Mølle) wasestablished in River Gudenaa (Fig. 1). Since 1987, the local author-ities (Vejle County and Horsens Municipality), along with the Na-tional Forest and Nature Agency, had debated with stakeholders forthe removal of the Vilholt dam to restore natural conditions andfaunapassage in the river. The dam was finally removed in 2008

65

after nearly two decades of debate. Lake Mossø is located approx-imately 6 km downstream of where the dam used to be. The riversystem is now home to a large population of brown trout (S. trutta),with Lake Mossø serving as highly productive feeding grounds forlake-dwelling brown trout (herein referred to as lake trout). Theselake trout originate from the spawning and nursery areas of RiverGudenaa, migrate down to the lake to feed, and return to the riverto spawn.

2.2. Electrofishing surveys

Starting in 1997 through 2016, electrofishing surveys wereconducted (end of August to beginning of October) 1.5 km upstreamof the damwithin the ponded zone (Fig. 1, A). Prior to removal, thedecreased velocity and increased water depth in this area led to theaccumulation of sand and silt on the bottom, with a minimumwater depth of approximately 0.7 m. Following the removal of thedam, the ponded zone disappeared and the natural shallow waterhabitat was restored to its original state, with faster-flowing water,a water depth of 10e30 cm, a natural substrate dominated bystones and cobbles, the original gradient (approx. 0.3%) and thepresence of water riffles, thus highly suitable brown trout (S. trutta)spawning and nursing grounds. It is worth noting that this type ofhabitat is scarce in larger Danish streams due to years of humanalterations, making this location of particularly high interest.

A second locationwas surveyed from 1987 through 2016, 1.5 kmdownstream of the dam (Fig. 1, B). This stretch was recognized asexcellent for spawning, even before the damwas removed. The laketrout from Lake Mossø gained access to this area after 1992, when afish ladder and a bypass stream were build at two weirs near thelake. Before 1992, the brown trout population was dependent onthe spawning of resident brown trout.

2.3. Fish density: mark-and-recapture

In the fall, the upstream (from 1997 to 2016) and the down-stream (from 1987 to 2016) locations were surveyed for lengths of160 m and 600 m, respectively. The width of the river at these lo-cations was approximately 20 m. Each location was electrofishedonce using two electrodes, with all captured brown trout marked(fin-clipped in this case). The following day, the same locationswere electrofished a second time. All previously marked fish (i.e.,recaptures) and unmarked fish (i.e., new captures) were counted.The numbers were then used to calculate fish density estimates.Fish below 14 cm were considered young of the year (YOY) whilelarger fish (above 14 cm) were pooled together and consideredolder fish (OLD). The two groups were distinguishable due to abimodal length distribution. The following formula was applied tocalculate density estimates of brown trout:

N ¼ ðM þ 1ÞðC þ 1ÞRþ 1

where, N is the density estimate, M is the number of fish caught andmarked during the first sampling, C is the total number of capturedfish during the second sampling (including recaptures), and R is thenumber of recaptures during the second sampling (Lockwood andSchneider, 2000). Results are presented as number of fish per me-ter (length) of river, in accordance to the national Danish BrownTrout Index (Kristensen et al., 2014), which states that populationestimates of YOY in Danish streams wider than 2 m should not becalculated as number per m2 as YOY mainly inhabit the river banks.

Fig. 1. The Vilholt dam was located in the Gudenaa river system, in central Jutland, Denmark, until 2008. The upstream and downstream sampling locations are represented byletters A and B, respectively.

K. Birnie-Gauvin et al. / Journal of Environmental Management 204 (2017) 467e471 469

2.4. Statistical analyses

Mann-Whitney U-tests were used to compare trout densitybefore and following removal of the Vilholt dam. The density (fishper m) of yearling (YOY) and older (OLD) fish were analyzedseparately in both the upstream (A) and downstream (B) zones. Theanalyses were done using R 3.1.2 (R Core Team, 2014). Variation inassociation with recorded mean values is given as standard devia-tion (±SD) throughout.

3. Results

An immediate increase of YOY brown trout was observed at theupstream stretch after removal of the dam, followed by a down-stream increase in YOY after three years. In the upstream zone,mean YOY density was 0.03 ± 0.04 fish per m before removal of thedam and 6.21 ± 2.77 fish per m following dam removal. The meanupstream OLD density before removal was 0.16 ± 0.08 fish per m,and 0.30 ± 0.07 fish per m following removal. The mean down-stream YOY density before and following the dam removal was1.2 ± 0.99 and 6.2 ± 2.8 fish per m, respectively. For OLD fish, themean downstream density was 0.31 ± 0.16 fish per m before damremoval and 0.43 ± 0.21 fish per m following dam removal.

In the upstream zone, both YOY (U ¼ 24.0, p ¼ 0.019) and OLDfish (U¼ 22.5, p¼ 0.041) densities increased significantly followingdam removal (Figs. 2A and 3A). In the downstream zone, YOYdensity increased significantly following dam removal (U ¼ 62,p < 0.001, Figs. 2B and 3), but no significant change in OLD density

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was found (U ¼ 46, p ¼ 0.14; Figs. 2B and 3B).

4. Discussion

The EUWater Framework Directive states that a watershed witha “good” ecological status should have biological elements thatshow little distortion as a result of anthropogenic activities, thoughthe quality of these elements may deviate slightly from thoseobserved in undisturbed conditions. A “high” ecological status re-quires that a system suffer no or very minor anthropogenic dis-turbances, with biological elements completely unaffected (EUWater Framework, 2000). Simultaneously, the European Renew-able Electricity Directive (2001/77/EC) encourages the use of small-scale hydropower facilities to generate renewable energy. Thepresence of dams (both small and large), and their associatedenvironmental and biological impacts to freshwater ecosystems,precludes an ecologically good status, as defined by the framework.The difficulty of achieving this status is further exacerbated by theencouragement of the directive to establish small hydrodams,making management and recovery plans contradictory and almostunachievable. Similar contradictive directives exist at the interna-tional level (e.g., Sustainable Development Goals by UnitedNations).

The availability and access to suitable habitats is of crucialimportance for a wide range of freshwater species, whether duringspawning migration, feeding or refuge seeking (Northcote, 1984;Taylor et al., 1993; Lucas and Baras, 2001). The observed increasein YOY density both upstream and downstream of where the

Fig. 2. Brown trout (Salmo trutta) density (number of individuals per m of river) up-stream (A) and downstream (B) of the Vilholt dam. Downward pointing arrow showsdam removal. Asterisks represent years when no surveys were carried out.

Fig. 3. Boxplots showing the density of YOY (A) and OLD (B) trout (Salmo trutta) in theupstream and downstream zones of the Vilholt dam before and after it was removed.The line within each box represents median fish density, ends of boxes represent the25th and 75th percentiles, and whiskers represent the 10th and 90th percentiles.Asterisks indicate significant difference at p < 0.05. Note the different scales on y-axes.

K. Birnie-Gauvin et al. / Journal of Environmental Management 204 (2017) 467e471470

Vilholt dam was located, along with the upstream increase in OLDfish, suggests that (1) the natural habitat quality was restored in theponded zone as a highly suitable spawning and nursing habitats,(2) safe passage and access to highly suitable spawning habitatupstream was reestablished, and (3) movement between the twospawning grounds increased recruitment. Here, we demonstratethat restoring river connectivity has allowed for a huge number offish to be born and thrive in an area previously devoid of YOY fish,presumably due to restored spawning habitat and the ease of accessto these high quality spawning grounds. The recorded density ofYOY trout in the present study (mean ¼ 6.2 YOY/m on bothstretches) place the river in “high ecological status” according tothe EU Water Framework Directives (the Danish threshold is 2.5YOY/m) and is in fact greater than normally observed in largeDanish rivers, suggesting that barrier removal may be the bestmitigation approach in the context of river restoration in frag-mented rivers.

The removal of the Vilholt dam restored the naturally adequatetrout habitat in the former ponded zone, resulting in an immediateincrease in both YOY and OLD fish upstream in 2009. This is likelybecause the removal allowed for the upstream passage of spawnersfrom the lake, along with providing highly suitable habitat foryoung fish to thrive, thus increasing survival. The removal had littlephysical effect on the downstream habitat, which was alreadysuitable for spawning. We note that beginning in 1992, an increasein OLD fish was observed downstream. This is due to the estab-lishment of a fish ladder and a bypass at twoweirs located betweenVilholt and LakeMossø. These structures led to a larger YOY densityin 1993. We also note that a sudden decrease in fish was observedin 1994; a large storm caused the dam to break down, letting largeamounts of mud and silt to be flushed downstream, practically

67

eliminating the year class. The year following removal (2009),neither YOY nor OLD fish densities increased downstream of thedam. In 2011, a large increase in YOY individuals downstream wasobserved. The large increase in YOY upstream in 2009 would haveyielded a large smolt cohort (length 12e15 cm) which likelymigrated down to Lake Mossø. These individuals would then bereturning to spawn in both stretches in winter 2010e2011, likelycontributing to the large YOY density observed in 2011 both up-stream and downstream of the former dam. Furthermore, it is alsopossible that YOY from upstream moved downstream to findsuitable habitat if the density of fish is too high upstream.

We have shown that barrier removal can be beneficial for fishdensity especially upstream, but also downstream. Since theremoval, local anglers have also noticed an increase in the size andnumber of lake trout caught in Lake Mossø. While these observa-tions suggest that the removal of an artificial obstacle may bebeneficial at a whole-system level, we cannot make that conclusionfor certain as our study did not specifically evaluate this. While theGudenaa river system supports a sustainable population of olderfish, including returning lake trout spawners from Lake Mossø aswell as resident trout, a wide spatial distribution of spawning andrecruitment is needed to maintain population levels over time(Berkeley et al., 2004). Before the Vilholt damwas removed, the rateof spawning was low, with few YOY surviving in the ponded zone.YOY are an important component for maintaining population

K. Birnie-Gauvin et al. / Journal of Environmental Management 204 (2017) 467e471 471

sustainability, and barriers may truncate the age-structure and therange of distribution of fish species, with potentially devastatingeffects on population sustainability.

This study demonstrates the extent to which small-scale ob-stacles (a 2.4 m high dam in this case) can affect the density anddistribution of river spawning fish. Low-head barriers of this type,which can obviously lead to the deterioration of natural spawningand nursery areas in ponded zones, are rarely considered in man-agement plans. It is our hope that these results will reinforce theneed to firstly, include smaller weirs and dams in managementplans, and secondly, considered removal as an option rather thanimmediately attempt to establish artificial fish passage. Our find-ings have important implications for the management of barriersacross the world. Environmental directives from many agencies(e.g., EUWaterframe Directive, UN Sustainable Development Goals)have made contradicting requests, with emphasis on reducingpollution, but little to no demands made to improve ecosystemsimpacted by barriers. Given the immediate positive effects of theremoval of small barriers, this approach should be viewed as aneconomically and ecologically profitable option.

Authors' contributions

KBG participated in the data analysis, data interpretation,manuscript conception and revision. JN participated in the dataacquisition and interpretation, as well as manuscript revision. KAand MHL participated in data interpretation and manuscriptrevision.

Data accessibility

Data available at https://doi.org/10.6084/m9.figshare.5387590.

Acknowledgments

We are thankful to all the volunteers (including communitymembers and anglers) who helped with the electrofishing surveysthrough the years. Funding for this research was provided by theDanish Rod and Net Fish License and the European Union AMBER(Adaptive Management of Barriers in European Rivers, Grantnumber 689682) project.

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Manuscript IV Catchment-scale effects of river fragmentation: a case study on restoring

connectivity

Under review in: Journal of Environmental Management

69

Catchment-scale effects of river fragmentation: a case study on restoring connectivity

Under review in Journal of Environmental Management

Kim Birnie-Gauvin1, Jan Nielsen1, Sten Bøgild Frandsen2, Hans-Martin Olsen3 and Kim

Aarestrup1

1 DTU Aqua, Section for Freshwater Fisheries and Ecology, Vejlsøvej 39, 8600 Silkeborg

2 Kolding Kommune, Akseltorv 1, 6000 Kolding

3 COWI A/S, Jens Chr. Schous Vej 9, 8000 Aarhus C

Author for correspondence: K. Birnie-Gauvin

kbir@aqua.dtu.dk

Running title: River fragmentation and restoration

Author contributions

KBG, JN and KA conceived and designed the approach. JN, SBF and HMO collected the

data. KBG collated and analysed the data. KBG and KA interpreted the data. KBG wrote the

first draft of the manuscript. KBG, JN, SBF, HMO and KA commented and edited the

manuscript.

70

Highlights

• We use a BACI approach to evaluate the effects of a pseudo dam removal that

reconnected one of two affected tributaries

• We found an immediate increase in YOY brown trout following the restoration

project

• Dam removal is an efficient method to restore river connectivity and habitat

suitability

Abstract

More than two thirds of large rivers worldwide are fragmented, threatening freshwater

biodiversity, river integrity, and the services that freshwater ecosystems provide for human

populations around the globe. These ecosystems are facing an unprecedented crisis, in large

part due to damming. In an effort to alleviate the impacts of barriers, engineered solutions

have been developed, though with somewhat underwhelming results. River restoration, and

dam removal especially, are viewed as the optimal option though seldom the go-to approach.

In this study, we evaluated the effects of a large restoration project (pseudo dam removal) in

River Kolding, Southern Jutland, Denmark, aimed at restoring connectivity to one of the

river’s main tributaries. The project came after the construction of a small bypass was

unsuccessful. We found that habitat connectivity was restored successfully in the tributary,

with a large increase in young-of-the-year brown trout (Salmo trutta) at reconnected sites,

reaching similar densities to downstream (non-affected) sites. We further observed a decrease

in length at reconnected sites, suggesting that natural spawning and rearing habitats were

successfully restored too. This is the first study that we know of using a BACI approach for

evaluating both the impacts of an engineered bypass and a ‘pseudo dam removal’ at a

catchment-scale, and highlights the need for prioritizing management and restoration at this

level.

Keywords: fish passage, fragmentation, hydropower, migration, restoration, Salmo trutta

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1. Introduction

Barriers are ubiquitous in today’s rivers, affecting the connectivity that is vital to the

ecological integrity of freshwater ecosystems (Bednarek 2001; Grill et al. 2015; Rincón et al.

2017). Worldwide, more than two thirds of large rivers are estimated to be fragmented (Grill

et al. 2015), with an estimated 2.8 million dams (with reservoir capacities greater than 103

m2, Lehner et al. 2011). Barriers fragment previously continuous populations, causing issues

with migration and fish passage, but also with critical spawning habitats and gene glow (Silva

et al. 2018; Birnie-Gauvin et al. 2019; Wilkes et al. in press). By altering hydrological

regimes and converting lotic habitat into lentic ones (Magilligan & Nislow 2005), dams

truncate the distribution and abundance of aquatic species, prevent fish movements, alter the

composition of aquatic communities by decreasing biodiversity, and increase the risk of

extinction (Lucas & Baras 2001; Ding et al. 2018).

Big or small, barriers pose several problems for fish on the move. Alterations of the

flow characteristics of the water affect important migration cues (Bunn & Arthington 2002).

Impoundments created by barriers cause delays in migration and act as sinks where fish are

highly susceptible to predation (Jepsen et al. 1998; Lucas & Baras 2001). An additional

consequence to barriers, often overlooked, is how much their presence alters the upstream

habitat, destroying spawning grounds critical for rheophilic species such as Atlantic salmon

(Salmo salar) and brown trout (S. trutta; Birnie-Gauvin et al. 2017a). Yet, maintaining

connectivity and effective ecosystem functioning in rivers is necessary to achieve a ‘good

ecological status’ according to the Water Framework Directive (Directive 2000/60/EC).

While large dams are often associated with large impacts, small barriers (with

reservoirs <0.1 km2) are present in much greater numbers, and make up approximately 99.5%

of the estimated number of artificial barriers globally (Lehner et al. 2011). Furthermore, the

number of barriers worldwide is likely to be underestimated; a recent study in the United

Kingdom found that barrier density was underestimated by 68%, particularly for small

barriers, and that more than 97% of rivers in the UK were fragmented (Jones et al. 2019). As

such, the consequences of barriers – worldwide – are likely to be highly underrated.

Despite removal being the only viable option for restoring natural habitats in their

entirety (Birnie-Gauvin et al. 2017a), engineered solutions are often the go-to tool when it

comes to barriers and fish passage (Birnie-Gauvin et al. 2019). Nonetheless, dam removal has

become an increasingly common approach for restoring rivers, especially as old barriers

become functionally obsolete or structurally unsafe (O’Connor et al. 2015).

72

Several studies have recently been published (O’Connor et al. 2015), showcasing

the positive effects of restoring connectivity, both in the short-term (e.g., Birnie-Gauvin

2018b; Ding et al. 2018) and the long-term (e.g., Catalano et al. 2007; Birnie-Gauvin et al.

2107b), but none that we know of have evaluated the effects on a catchment-scale. Much of

the research has rather focused on the local (site-specific) or river-scale effects. Hence, the

aim of this study was to evaluate the large-scale effects of removing a 26 m hydrodam from a

tributary to a major river in Denmark on the density of brown trout. This case study is

particularly interesting because the dam originally affected two tributaries, but an artificial

stretch of river was constructed for the largest tributary (river Vester Nebel) to restore it to its

natural state. This restoration project is somewhat unusual as the dam was (in essence)

removed (referred to as pseudo dam removal herein) from one tributary (river Vester Nebel)

and maintained in another (river Almind), but it allows us to investigate the wider catchment-

scale impacts of river fragmentation and connectivity reestablishment using a before-after-

control-impact (BACI) approach.

2. Methods

2.1 Study site

In 1920, the Harte hydropower plant was established near river Kolding, southern Jutland,

Denmark (Figure 1). The dam was in fact built in an artificial channel (red line in Figure 1)

– dug specifically for this purpose – that would allow for a greater drop (and thus a larger

energy production) than if it had been built in the natural river. In an effort to divert as much

water toward the dam as possible, a diversion ‘wall’ was also built at the confluence of river

Almind and river Vester Nebel (two tributaries to river Kolding) to force water from both

tributaries to flow toward the dam, and thereby blocking passage to the upper reaches of both

river Almind and Vester Nebel (white square in Figure 1). The dam was (and still is) the

highest in Denmark with a height of 26 m. At the time of establishment, the hydropower plant

produced half of the electricity used in the catchment area, but the growing demand for

electricity implied that the production around year 2000 was limited to the demands from

only 450-500 households.

With the growing realization of the negative effects of the dam on the local aquatic

fauna (brown trout Salmo trutta in particular) however – incompatible with the environmental

goals of the local community – a decision was made in 1993 to make a small bypass (150 l s-

1) around the diversion wall to re-establish connectivity in river Vester Nebel. Unfortunately,

73

this effort did not have the intended effects, as the habitat upstream of the dam was still poor,

up- and downstream passage was inadequate, and very few adult spawners made it upstream

of the bypass to spawn (S. B. Frandsen, personal communication). Fish were still required to

move through impoundments, and experienced high predation. After several years of

negotiations between the local communities, the Forest and Nature Agency, the sports fishing

associations and the hydropower plant owner, it was decided in 2007 that river Vester Nebel

would be returned to its natural state, and the power plant would rely solely on water intake

from river Almind (approx. 1/3 of its previous water intake).

2.2 The restoration project: pseudo dam removal

In the spring of 2008, a new stretch of 1025 m of river was created around the diversion wall

to reconnect the upper reaches of river Vester Nebel to the mainstem of river Kolding

(Figure 1). Importantly, the new stretch has a total drop of 4.5m (equivalent to 4.39m drop

per km) which reflects the naturally low gradient found across Denmark. The new stretch was

opened in June 2008. Now, passage to the upstream reaches of river Vester Nebel was

completely unhindered, freeing 41 km of river and a total catchment area of 89 km2. Fish no

longer needed to overcome impoundments during their up- and downstream migration, and

the natural spawning grounds were again accessible. Thus, the hydropower plant was

‘removed’ from river Vester Nebel in June 2008. A total of 35 000 m3 of soil was moved to

create the new river stretch, and 2 500 m3 of stone, boulders and spawning gravel was added.

The total cost was approx. 4.2 million Danish kroners (equivalent to 560 000 euros). The

project restored free access to more than 40 km of free river (considered large in Denmark,

where most rivers are <15 km).

74

Figu

re 1

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75

2.3 Electrofishing sites and surveys

Electrofishing surveys for brown trout were performed as part of the Kolding river

management plans in 1992, 2001, 2008 and 2017 at 74 different sites. The sites were divided

into three types: 1) downstream sites (n = 46), which are sites that were never directly

affected by the dam (i.e., always below the dam), 2) reconnected sites (n = 21), which are

sites that were above the dam in river Vester Nebel (i.e., regained connectivity following the

restoration project), and 3) regulated sites (n = 8), which are sites that remain affected by the

dam in river Almind (i.e., always above the dam). We note that two of the downstream sites

in river Vester Nebel (those in the main stem downstream of the diversion wall, #44 and 45)

would have had a reduced water flow due to the diversion wall. These are nonetheless

included in the downstream sites group because it would be difficult to quantify the effects of

the diversion wall on only two sites. Note that not all sites were surveyed each year (the

lowest number of surveys was in 1992 with 30 downstream sites, 14 reconnected sites and 5

control sites; see Table 1 for details). Furthermore, to investigate the immediate effects of the

pseudo dam removal, electrofishing surveys were conducted at 12 reconnected sites in 2009

(1 year after the restoration). For detailed location of each site, see Appendix.

To evaluate fish density, the two-pass depletion method for density surveys was

used (Lockwood and Schneider 2000), and followed this formula:

𝑁𝑁 = 𝑦𝑦12

𝑦𝑦1 − 𝑦𝑦2×

1𝐴𝐴

× 100𝑚𝑚2

where N is the density estimate per 100m2, y1 is the number of fish caught during the first

electrofishing pass, y2 is the number of fish caught during the second electrofishing pass and

A is the area surveyed in m2.

All captured fish were measured (total length, ±0.5cm). Based on the length

distribution of fish at each site, we were able to classify fish into two groups: 1) young-of-

the-year (YOY, 0+) and 2) older year classes (OLD, 1+). This grouping prevents effects from

confounding factors such as variation in sampling time and water temperature (where growth

rate could be higher or lower with higher or lower temperature, respectively).

We note that no stockings of brown trout were made prior to the electrofishing in

1992 and 2001, and since 2007 all stockings of brown trout in the watershed have been

suspended. Consequently, YOY caught at the electrofishing were produced by spawning sea

trout as well as all OLD brown trout since 2007.

76

2.4 Statistical analyses

Given the variability in fish densities at stations within the same site type, we opted to use a

generalized linear model (GLM) with negative binomial distribution for count data (Zuur et

al. 2013) and log link function. To obtain the total count at each station, we added the number

of fish captured at the first and second electrofishing pass. To account for variations in the

area surveyed, our model included an offset for area (in m2). To test for immediate restoration

effects (between 2008 and 2009) on a local scale, our model was as follows:

𝐹𝐹𝐹𝐹𝐹𝐹ℎ 𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝑖𝑖 ~ 𝑁𝑁𝑁𝑁(𝜇𝜇𝑖𝑖 ,𝑘𝑘)

𝐸𝐸(𝐹𝐹𝐹𝐹𝐹𝐹ℎ 𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝑖𝑖) = 𝜇𝜇𝑖𝑖

𝑣𝑣𝑣𝑣𝑣𝑣(𝐹𝐹𝐹𝐹𝐹𝐹ℎ 𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝑖𝑖) = 𝜇𝜇𝑖𝑖 + 𝜇𝜇𝑖𝑖2

𝑘𝑘log(𝜇𝜇𝑖𝑖) = 𝛾𝛾𝑖𝑖

𝛾𝛾𝑖𝑖 = 𝛼𝛼 + 𝛽𝛽1 ∙ 𝑌𝑌𝑌𝑌𝑣𝑣𝑣𝑣𝑖𝑖 + 𝐶𝐶𝑜𝑜𝑜𝑜𝐹𝐹𝑌𝑌𝐶𝐶(𝐴𝐴𝑣𝑣𝑌𝑌𝑣𝑣)

Where α is the intercept, and β is the slope. To test for more long term effects, we tested for

effects of site type, year and their interaction on fish count across years. Our model was as

follows:

𝛾𝛾𝑖𝑖 = 𝛼𝛼 + 𝛽𝛽1 ∙ 𝑌𝑌𝑌𝑌𝑣𝑣𝑣𝑣𝑖𝑖 + 𝛽𝛽2 ∙ 𝑇𝑇𝑦𝑦𝑇𝑇𝑌𝑌𝑖𝑖 + 𝛽𝛽3 ∙ 𝑌𝑌𝑌𝑌𝑣𝑣𝑣𝑣𝑖𝑖 ∙ 𝑇𝑇𝑦𝑦𝑇𝑇𝑌𝑌𝑖𝑖 + 𝐶𝐶𝑜𝑜𝑜𝑜𝐹𝐹𝑌𝑌𝐶𝐶(𝐴𝐴𝑣𝑣𝑌𝑌𝑣𝑣)

Where α is the intercept, and β is the slope. Note that the same model was performed to

analyse YOY and OLD data separately.

We investigated differences in average length across site types and year by

performing a GLM with Gaussian distribution. Length data for OLD fish was ‘log + 1’

transformed to meet underlying normality assumptions. Site type, year and their interaction

were included as explanatory variables. Our model was as follows:

𝐿𝐿𝑌𝑌𝐶𝐶𝐿𝐿𝐶𝐶ℎ𝑖𝑖 = 𝛼𝛼 + 𝛽𝛽1 ∙ 𝑌𝑌𝑌𝑌𝑣𝑣𝑣𝑣𝑖𝑖 + 𝛽𝛽2 ∙ 𝑇𝑇𝑦𝑦𝑇𝑇𝑌𝑌𝑖𝑖 + 𝛽𝛽3 ∙ 𝑌𝑌𝑌𝑌𝑣𝑣𝑣𝑣𝑖𝑖 ∙ 𝑇𝑇𝑦𝑦𝑇𝑇𝑌𝑌𝑖𝑖 + 𝜀𝜀𝑖𝑖

𝜀𝜀𝑖𝑖 ~ 𝑁𝑁(0,𝜎𝜎2)

Where α is the intercept, β is the slope and ɛi is normally distributed noise with mean 0 and

variance σ2. Note that the same model was performed to analyse YOY and OLD data

separately. Statistical significance was evaluated with likelihood ratio tests (LRT) with p <

0.05. All statistical analyses were performed in R version 3.6.1 (R Core Team 2018).

77

3. Results

Density of YOY varied considerably over the years and between stations (Table 1).

Generally, we note an increase in YOY density at downstream sites, and an increase in 2017

at reconnected sites where the pseudo removal of the dam had re-established connectivity in

River Vester Nebel. The opposite trend was observed at regulated sites that remain affected

by the dam in river Almind, with most sites undergoing a decrease in YOY density. Density

of OLD varied less, and tended to be fairly constant over the years.

YOY count immediately increased significantly (by a factor of 4) at reconnected

sites after the restoration project (LRT = 12.689, df = 1, p = 0.0004, Figure 2), though OLD

count did not change as a result of the restoration (LRT = 1.0998, df = 1, p = 0.29, Figure 3).

Year and site type significantly affected the number of YOY (year × type, LRT =

33.1, df = 6, p < 0.0001, Figure 4) but not OLD fish (all p > 0.05, Figure 5). More

specifically, YOY counts in the regulated sites were significantly lower in 2017 (post-

restoration) than in the downstream and reconnected sites (p = 0.0003). YOY counts in the

reconnected sites in 2008 were lower than in the downstream sites (p = 0.001), as were those

in the regulated sites though only marginally significantly (p = 0.05). We note that there is

more variation in the fish counts at regulated sites than the other site types, likely exacerbated

by the low site numbers.

Average length of YOY fish was also influenced by year and site type (year × type,

LRT = 851.3, df = 6, p < 0.0001, Figure 6). Specifically, the average length of YOY in 2008

was significantly higher at regulated and reconnected sites than at downstream sites (p <

0.0001). Average length of YOY was also higher at reconnected sites than downstream sites

in 2001 (p < 0.0001). The average length of YOY was significantly lower in the downstream

and reconnected sites than in the regulated sites in 2017 (p < 0.001), where a sharp decrease

in average length between 2008 and 2017 was particularly obvious in the reconnected sites.

Similarly, the average length of OLD fish was also affected by year and type (year

× type, LRT = 191.8, df = 6, p < 0.0001, Figure 7). In 2001, the average length of OLD fish

at downstream sites was shorter than at regulated (p = 0.004) and reconnected sites (p <

0.0001). The average length of OLD fish at downstream sites was also shorter than regulated

(p < 0.0001) and restored sites (p < 0.0001) in 2008. In 2017, the average length of OLD fish

was longer at regulated sites than both downstream and reconnected sites (p = 0.0003).

78

Table 1. Density estimates per 100m2 of young-of-the-year (YOY) and older year classes

(OLD) at each station according to type and year. See Appendix for details on site location. Station Type YOY OLD

1992 2001 2008 2017 1992 2001 2008 2017

1 Downstream 6.7 63.6 17.0 0.0 0.0 5.2 1.5 0.0

2 Downstream 5.6 92.9 37.0 19.4 4.3 9.5 8.3 3.2

3 Downstream 54.6 39.0 7.6 102.5 1.3 18.0 12.4 16.0

4 Downstream 19.1 48.8 45.3 94.5 30.5 72.2 30.1 25.2

5 Downstream NA 107.3 49.2 106.1 NA 39.1 21.6 20.4

6 Downstream 71.1 14.8 28.6 66.5 25.2 34.6 18.3 33.0

7 Downstream 86.7 16.0 NA 192.4 20.0 26.7 NA 31.8

8 Downstream 27.2 28.1 NA 109.2 13.3 3.1 NA NA

10a Downstream 62.6 NA 50.0 198.9 8.7 NA 16.7 0.0

10b Downstream 16.1 NA NA 4.0 53.4 NA NA 0.0

11 Downstream NA 0.0 0.0 0.0 NA 0.0 0.0 0.0

12 Downstream 1.0 36.5 1.3 7.8 1.0 3.3 3.1 2.8

13a Downstream NA NA 9.4 46.7 NA NA 12.5 25.0

14a Downstream NA NA 264.0 270.7 NA NA 14.3 0.0

14b Downstream NA NA 159.1 138.9 NA NA 9.1 0.0

14 Downstream 9.9 3.8 58.3 52.6 13.4 3.8 17.3 16.3

16 Downstream NA 11.0 4.5 169.9 NA 3.8 2.7 0.0

17 Downstream 5.6 82.0 0.0 127.3 2.2 4.4 3.1 21.4

18 Downstream 0.0 5.0 2.9 150.4 0.0 0.0 1.4 7.5

19 Downstream NA 22.2 1.2 128.5 NA 0.0 10.8 9.0

20 Downstream 7.1 71.5 38.7 144.5 0.0 6.3 12.8 20.0

21 Downstream 6.7 33.3 5.4 48.0 0.0 14.8 7.0 6.8

22 Downstream 63.8 85.9 46.6 113.3 33.7 22.3 9.9 13.6

23 Downstream 15.6 0.0 54.4 56.8 15.6 0.0 16.1 23.0

24 Downstream NA 116.8 66.9 141.8 NA 7.7 3.7 0.0

25 Downstream 412.0 96.8 155.2 159.1 52.1 20.0 0.0 0.0

26a Downstream NA NA 19.4 14.9 NA NA 29.0 0.0

26 Downstream NA 0.0 0.0 0.0 NA 0.0 0.0 0.0

27 Downstream NA 50.9 11.7 128.4 NA 2.0 10.0 0.0

28 Downstream 12.6 75.1 271.0 104.5 14.1 46.6 61.6 0.0

29 Downstream NA 100.1 211.5 35.4 NA 7.3 28.9 0.0

30 Downstream 0.0 112.4 262.2 12.5 0.0 0.0 4.0 0.0

31 Downstream 42.5 74.5 120.4 286.5 0.0 7.7 14.3 3.1

32 Downstream NA 173.0 44.2 68.7 NA 0.0 0.0 0.0

33 Downstream 18.2 311.5 137.1 64.5 0.0 0.0 22.2 28.3

34 Downstream 3.1 235.0 337.0 67.2 0.0 3.2 61.0 44.5

35a Downstream NA NA 97.5 247.9 NA NA 1.3 0.0

35 Downstream 8.3 243.6 168.9 238.8 47.2 59.3 50.0 14.7

36 Reconnected NA 0.0 1.7 106.8 NA 7.1 0.0 0.0

37 Reconnected NA 48.5 13.5 135.9 NA 14.3 12.7 15.5

79

38 Reconnected 39.0 21.5 2.7 88.4 9.9 9.7 4.9 25.2

39 Reconnected 50.0 56.3 7.9 99.9 35.8 13.9 6.4 54.1

40 Reconnected 23.3 13.6 10.3 188.4 3.3 17.8 6.8 5.8

41 Reconnected 25.7 10.1 4.4 231.2 15.4 24.2 3.1 12.0

42 Reconnected 4.8 11.6 28.2 167.0 6.1 18.5 9.9 48.6

43 Reconnected 3.3 26.3 9.9 158.1 9.6 22.0 5.5 25.2

44 Downstream 20.6 50.4 66.0 89.8 9.4 47.1 NA 23.3

45 Downstream 3.2 39.3 NA 101.3 5.8 25.3 NA 19.3

47 Reconnected NA 34.0 38.9 188.2 NA 5.3 8.2 0.0

48 Reconnected 30.3 40.9 0.9 131.9 6.4 9.1 3.5 0.0

49 Reconnected NA 38.7 NA 38.2 NA 1.4 NA 0.0

50 Reconnected 8.0 134.7 7.3 50.1 0.0 6.7 2.7 0.0

51 Reconnected 95.5 107.0 27.8 NA 3.7 17.0 5.6 0.0

52 Reconnected 58.7 94.1 9.1 68.7 86.0 21.4 8.9 0.0

53 Reconnected 74.6 100.0 88.8 38.3 44.0 17.5 6.7 15.5

54 Reconnected 28.3 35.6 34.4 155.3 6.7 24.4 26.8 0.0

55 Reconnected 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0

56 Reconnected NA 1.7 21.6 39.2 NA 6.7 12.9 0.0

57 Reconnected NA 33.3 6.0 174.2 NA 8.3 8.0 8.6

58 Downstream NA NA 99.6 73.7 NA NA 26.0 0.0

59 Downstream NA 72.9 58.7 87.8 NA 21.4 7.8 0.0

61 Downstream 0.0 224.7 172.5 156.5 0.0 56.0 52.6 10.4

62 Downstream 0.0 213.8 140.3 257.0 0.0 40.6 27.2 44.6

63 Downstream 98.5 238.0 236.4 255.8 22.2 68.8 39.3 23.2

64 Downstream 0.0 40.5 307.3 93.1 0.0 4.0 23.3 0.0

65 Reconnected NA 81.3 40.1 185.3 NA 2.2 1.4 0.0

66a Reconnected NA NA 31.4 396.0 NA NA 0.0 3.1

66 Reconnected 0.0 4.0 NA 0.0 0.0 0.0 NA 0.0

67 Regulated 131.5 108.5 79.5 22.9 64.5 64.3 31.3 1.1

68a Regulated NA NA 63.1 27.5 NA NA 13.7 51.2

68b Regulated NA NA 9.2 NA NA NA 6.2 NA

68 Regulated 78.4 188.9 NA 46.4 31.8 12.5 NA 1.2

69 Regulated 0.0 1.5 0.0 0.0 0.0 0.0 0.0 0.0

70 Regulated 0.0 10.9 59.8 0.0 25.0 16.4 4.4 0.0

71 Regulated 32.0 13.3 32.5 0.0 14.0 13.0 23.4 0.0

72 Regulated NA 10.0 11.7 0.0 NA 5.0 11.7 0.0

80

Figure 2. Immediate effects of restoration. Modelled density of brown trout (Salmo trutta)

young-of-the-year (YOY) per 100m2 of river before (2008) and after the pseudo-removal

(2009) at 12 reconnected sites (with 95% confidence intervals). A significant increase was

observed (p < 0.0004).

81

Figure 3. Immediate effects of restoration. Modelled density of brown trout (Salmo trutta)

older year classes (OLD) per 100m2 of river before (2008) and after the pseudo-removal

(2009) at 12 reconnected sites (with 95% confidence intervals). No significant differences are

present.

82

Figure 4. Modelled density of brown trout (Salmo trutta) young-of-the-year (YOY) per

100m2 of river before the small bypass (1992), before the pseudo-removal (2001, 2008) and

after the pseudo-removal (2017) at downstream, regulated and reconnected sites (with 95%

confidence intervals). A significant year × type interaction is present (p < 0.0001).

83

Figure 5. Modelled density of brown trout (Salmo trutta) older year classes (OLD) per

100m2 of river before the small bypass (1992), before the pseudo-removal (2001, 2008) and

after the pseudo-removal (2017) at downstream, regulated and reconnected sites (with 95%

confidence intervals). No significant differences are present.

84

Figure 6. Average length of brown trout (Salmo trutta) young-of-the-year (YOY) before the

small bypass (1992), before the pseudo-removal (2001, 2008) and after the pseudo-removal

(2017) at downstream, regulated and reconnected sites (with 95% confidence intervals). A

significant year × type interaction is present (p < 0.0001).

85

Figure 7. Average length of brown trout (Salmo trutta) older year classes (OLD) before the

small bypass (1992), before the pseudo-removal (2001, 2008) and after the pseudo-removal

(2017) at downstream, regulated and reconnected sites (with 95% confidence intervals). A

significant year × type interaction is present (p < 0.0001).

86

4. Discussion

Habitat connectivity is a necessary component of healthy freshwater communities and should

thus be a conservation priority (Abell et al. 2011; Reid et al. 2018). Our study supports the

importance of access to upstream habitats, and demonstrates that restoring connectivity at a

particular site can be highly beneficial for fish populations. Our findings also highlight that

catchment-scale management and restoration should be prioritized; while restoration at one

location is a step in the right direction, leaving other sites fragmented and impacted by

anthropogenic structures (i.e. regulated sites in this study) will always have the opposite

effect.

We documented an immediate increase in YOY at reconnected sites following the

pseudo-removal, similar to the densities at downstream sites, suggesting that the restoration

project successfully re-established safe passage and connectivity in River Vester Nebel.

The absence of differences in YOY density between 1992 and 2001 at reconnected sites also

indicates that the initial bypass constructed around the diversion wall failed to restore

connectivity, and indicates that this option is suboptimal for restoring connectivity. This is

further supported by the observation that very few adult spawners made it upstream of the

bypass (S. B. Frandsen, personal communication). YOY densities at regulated sites steadily

decreased between 1992 and 2017, perhaps indicating a worsening of environmental

conditions or decreased recruitment at those sites due to inaccessibility. Adult sea trout likely

opted to spawn in locations with more suitable spawning habitats (i.e., gravel substrate,

highly oxygenated water) and higher flow (Armstrong et al. 2003), such as in the reconnected

river Vester Nebel, rather than to overcome the Harte hydropower. We also noted a steady

increase in YOY at downstream sites, likely because these sites were subjected to other

management interventions from the local municipality, such as small restoration projects to

improve habitats. The local municipality of Kolding is known (in Denmark) to have an active

stream restoration policy. Concurrently, the increase in YOY density could, at least in part, be

due to the increased abundance of and access to suitable spawning habitat created by the

restoration project (Birnie-Gauvin et al. 2017b). Spawning and recruitment over a wide

spatial distribution is essential to maintaining population sustainability over time (Berkeley et

al. 2004).

The density of OLD fish remained constant over the years at all site types. While

density-dependent factors tend to regulate population abundance in the early stages, density-

independent factors tend to determine abundance thereafter (Milner et al. 2003). As such, the

density of OLD fish should be determined by factors such as temperature, oxygen, substrate,

87

and other water quality parameters (reviewed in Milner et al. 2003). However, there is also

evidence suggesting that intraspecific competition between cohorts (YOY vs OLD in this

case) can affect the density of an age class (Elliott 1994; Nordwall et al. 2001). Thus, it is

possible that despite the improvement of conditions (density-independent factors) at

reconnected sites, which should have yielded an increase in OLD fish, the large increase in

YOY led to increased competition between the year classes, and perhaps led to the

displacement of OLD fish to other unsampled areas. Alternatively, fish may migrate out at a

younger age/smaller size also due to increased competition. This is corroborated by the

decrease in average length of OLD fish between 2008 and 2017 at reconnected sites. At

downstream and regulated sites, habitat conditions and YOY densities were more constant

over the years, and thus the density of OLD fish remained equally constant.

The decrease in length of YOY observed between 2008 and 2017 at reconnected

sites, similar to that of downstream sites, might be the result of a few factors: 1) more adult

spawners due to free access to upper reaches of the river, and hence an increase in egg

deposition (i.e., overall increased abundance of fish with varying sizes); 2) greater survival of

fry due to adequate rearing habitat, and thus a greater number of smaller fish overall; 3) the

passage of smaller fish, which was likely more hindered by the barrier than that of larger fish,

is no longer affected, and thus these fish can move further upstream with no restriction; or 4)

the overall increase in abundance of fish has caused an increase in competition and thus fish

are smaller in size, or migrate at a younger age/smaller size (i.e. density-dependent

regulation, Jenkins et al. 1999; Milner et al. 2003). A combination of all four factors may be

at play, but the increased habitat suitability for spawning and early growth, and resulting

increased survival and competition is likely to be the most important elements at play. A

decrease in YOY average length was also observed at regulated sites between 2008 and 2017,

but this was based on so few fish (n = 125) in comparison to reconnected (n = 1848) and

downstream (n = 3656) sites that this result should be interpreted with caution. It is likely that

YOY were on average shorter because of worsening conditions of the habitat and availability

of food items. This is supported by the decrease in YOY density at regulated sites.

The average length of both YOY and OLD fish was more constant – though still

decreased – at downstream sites than reconnected and regulated. This is perhaps because

environmental conditions (including habitat suitability and food availability) were more

constant at these sites, and less susceptible to large fluctuations in abiotic factors like

temperature and flow, which tend to change drastically as a result of damming (Bednarek

2001), or due to a more steady number of adult spawners. The slight decrease in average

88

length may reflect the benefits of local restoration projects led by the local municipality,

which would increase the number of spawners, and enhance YOY survival overall, thus

indirectly increased competition within/across year classes (Nordwall et al. 2001).

We observed some variability across years within each site type, reflecting the true

nature of ecosystems. Variation in year classes due to abiotic factors like flow is not

uncommon in salmonids (Warren et al. 2015; Lobón-Cerviá et al. 2018), and could not be

controlled and/or accounted for in this study. Nonetheless, our findings clearly demonstrate

the benefits of barrier removal for salmonids. As opposed to some evidence suggesting that

fish communities may take decades to be reconnected (e.g. Poulos and Chernoff 2017), YOY

density responded immediately to this restoration project with an increase, similar to previous

work in similar rivers (Birnie-Gauvin et al. 2017b, 2018). This rapid increase is perhaps

possible because a large population of trout is present in the Kolding system, many of which

could reconquer river Vester Nebel. This is further corroborated by a large influx of sea trout

observed by local recreational anglers starting in 2012, which would correspond to trout born

in 2009 having spent 1-2 years in freshwater and 1-2 years at sea (S. B. F., personal

communication). Given the variability and flexibility in life history traits of brown trout, it is

also possible that this species responds faster to habitat restoration than other species (e.g.

Marttila et al. 2019).

Society is, in large part, failing its freshwater ecosystems. In Europe, the Water

Framework Directive (WFD), widely accepted as the most ambitious piece of European

environmental legislation, has yet to deliver its main objectives of non-deterioration of water

systems and achievement of good ecological status in all EU waters (Voulvoulis et al. 2017).

Fifteen years after its implementation, the WFD has gone from great expectations to

confusion, exemptions and time delays. While the WFD demands for no or very minor

human disturbances, the EU Renewable Electricity Directive (Directive 2001/77/EC) advises

for more small-scale hydropower, paving way for confusion among sectors. Any ‘heavily-

modified’ river is essentially exempt from abiding to the demands of the WFD (Directive

2000/60/EC). As such, rather than demanding for improvements in those poorly maintained

systems, the directive is effectively giving up on improving those systems. Even in rivers not

considered ‘heavily-modified’, like River Almind, very little action is done to improve their

status despite the WFD requiring action. We would argue that, at least for rivers altered by

barriers, like River Almind, dam removal should be strongly encouraged and implemented as

a tool to improve the conditions of those rivers.

89

4.1 Conclusions

The availability and access to adequate rearing, feeding and spawning habitats is a necessity

for most freshwater species, especially rheophilic ones like trout and salmon (Northcote

1984; Lucas & Baras 2001). Our study highlights the merits, and the need, for restoring

natural and connected freshwater habitats. Anthropogenic disturbance, and river

fragmentation in particular, is a major source of freshwater biodiversity loss worldwide,

contributing to severe population declines (Dudgeon et al. 2006; Reid et al. 2018). With

thousands of dams planned or under construction (Zarfl et al. 2015), thousands of obsolete,

unsafe and/or undocumented smaller barriers (Jones et al. 2019), and little consideration for

their ecological impacts (Winemiller et al. 2016), freshwater ecosystems are facing a major

crisis. While they comprise only 0.01% of the water on Earth, freshwater ecosystems host

approximately 9.5% of all known animal species (Balian et al. 2008). Free-flowing rivers are

a major component of these ecosystems, and increase their resilience against environmental

change including climate and land use changes (Groves et al. 2012; Grill et al. 2015). We

recommend dam removal become a go-to tool for addressing this crisis.

5. Acknowledgments

This study was funded by the Kolding Kommune and the Danish Net and Fish Licence

Funds.

6. Competing Interests

The authors declare no competing interests.

90

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Appendix

A1. Site locations. Red shaded area represents previously affected regions of the catchment by the

Harte hydropower dam.

96

Manuscript V River connectivity reestablished: effects and implications of six weir removals

on brown trout smolt migration

Published in: River Research and Applications, 34, 548-554 doi: 10.1002/rra.3271

97

Received: 22 November 2017 Revised: 7 March 2018 Accepted: 8 March 2018

DOI: 10.1002/rra.3271

R E S E A R CH AR T I C L E

River connectivity reestablished: Effects and implications of sixweir removals on brown trout smolt migration

K. Birnie‐Gauvin1 | M.M. Candee1,2 | H. Baktoft1 | M.H. Larsen1,3 | A. Koed1 |

K. Aarestrup1

1DTU AQUA, National Institute of Aquatic

Resources, Section for Freshwater Fisheries

Ecology, Technical University of Denmark,

Silkeborg, Denmark

2Freshwater Biology, University of

Copenhagen, Copenhagen, Denmark

3Danish Centre for Wild Salmon, Randers,

Denmark

Correspondence

K. Birnie‐Gauvin, DTU AQUA, National

Institute of Aquatic Resources, Section for

Freshwater Fisheries Ecology, Technical

University of Denmark, Vejlsøvej 39, 8600

Silkeborg, Denmark.

Email: kbir@aqua.dtu.dk

Funding information

AMBER (European Union), Grant/Award

Number: 689682; Danish Rod and Net Fish

License Funds

- - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - -

This is an open access article under the terms of th

the original work is properly cited.

© 2018 The Authors River Research and Applicati

River Res Applic. 2018;1–7.

Abstract

Today's river systems have been extensively modified, requiring us to rethink how we

approach the management of these important ecosystems. We evaluated the effects

of removing 6 weirs in River Villestrup (Jutland, Denmark) on the smolt run of brown

trout (Salmo trutta) over the course of 12 years. During 5 of these years, we evaluated

the number, size, and timing of smolts during their downstream migration. We found

an increase in smolt output following the weir removals, along with a decrease in

average length and indications of an earlier peak migration. Our results suggest that

barrier removal has led to an increase in spawning success by adults, fry survival,

recruitment, and smolt migration success. Weir removal is therefore a viable manage-

ment approach to restore connectivity in freshwater streams and rivers, which pro-

motes the passage of smolts as they migrate to marine environments.

KEYWORDS

barriers, freshwater ecosystems, migration, removal, river restoration, Salmo trutta, smolt

1 | INTRODUCTION

1.1 | River connectivity

The diversity, abundance, and sustainability of aquatic species have

long been threatened by the human‐induced fragmentation of rivers

(Khan & Colbo, 2008; Saunders, Hobbs, & Margules, 1991). Barriers

in the form of dams, weirs, and culverts have become so prominent

in today's river systems that the majority of them have lost their orig-

inal connectivity and natural characteristics (Jager, Chandler, Lepla, &

Van Winkle, 2001; Jungwirth, Schmutz, & Weiss, 1998). These barriers

exacerbate the current poor state of many freshwater ecosystems.

Efforts to mitigate the impacts of barriers, such as fishpasses, have

seen limited success (Bunt, Castro‐Santos, & Haro, 2012) and are usu-

ally costly (Gibson, Haedrich, & Wernerheim, 2005). Furthermore,

such approaches do not repair the damage done to the ecosystems

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e Creative Commons Attribution Li

ons Published by John Wiley & So

98

as a whole (Birnie‐Gauvin, Aarestrup, Riis, Jepsen, & Koed, 2017);

rather, they provide an opportunity for some fish to move upstream

or downstream past the barrier. This is particularly relevant for migra-

tory fish species such as salmonids, which depend on freshwater

migrations to complete their lifecycle (Jonsson & Jonsson, 1993;

Klemetsen et al., 2003). Better management tools need to be imple-

mented to promote the persistence of these migratory species, such

as barrier removal and other types of restoration projects.

1.2 | Brown trout

The brown trout (Salmo trutta, Salmonidae) is a partially anadromous

salmonid species, native to many regions of Europe (Jonsson &

Jonsson, 1993). Brown trout spawn in the upper reaches of rivers,

where the substrate is typically suitable for spawning and early

growth, and predators are typically absent (Armstrong, Kemp,

- - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - -

cense, which permits use, distribution and reproduction in any medium, provided

ns Ltd

wileyonlinelibrary.com/journal/rra 1

2 BIRNIE‐GAUVIN ET AL.

Kennedy, Ladle, & Milner, 2003; Shirvell & Dungey, 1983). Juvenile

trout generally spend between 1 and 5 years in freshwater, after

which individuals differentiate phenotypically (Nielsen, Aarestrup,

Nørum, & Madsen, 2003). Some individuals will assume a resident

phenotype and remain in freshwater their entire life, whereas others

will assume the migratory phenotype and migrate to marine environ-

ments (Jonsson & Jonsson, 1993; Nielsen, Aarestrup, & Madsen,

2006). This phenomenon is known as partial migration (Chapman,

Brönmark, Nilsson, & Hansson, 2011).

Although the drivers for partial migration remain poorly under-

stood (though many hypotheses exist, Chapman et al., 2011), the bene-

fits of migrating to sea appear to be linked to a larger availability of food

items in marine environments, thus allowing migratory individuals to

attain larger sizes and a greater reproductive potential (Chapman

et al., 2011; Northcote, 1984; Shrimpton, 2013). Juveniles that become

migratory individuals are known as smolts and differ from their resident

counterparts both behaviorally and physiologically. For example, smolts

appear to be less aggressive (Jonsson & Jonsson, 2011; Thorstad et al.,

2012), have greater sodium‐potassium ATPase activity in their gills

(Aarestrup, Nielsen, &Madsen, 2000), and appear to have greater levels

of blood‐circulating antioxidants (Birnie‐Gauvin et al., 2017).

Smolts typically migrate during the months of March to May

depending on latitude (peak smolt migration period, e.g., Bohlin,

Dellefors, & Faremo, 1993), though some migrate during the autumn

(Winter et al., 2016 Aarestrup, Birnie‐Gauvin, & Larsen, 2018). The

downstream smolt migration is thought to be triggered by a range of

environmental factors, such as photoperiod, temperature, and dis-

charge (Hoar, 1988). Furthermore, smolts are thought to migrate

downstream during the “smolt window.” This window is thought to

be affected by factors such as physiological and ecological readiness

to enter marine environments, risk of predation, and growth potential

(McCormick, Hansen, Quinn, & Saunders, 1998). It is thus essential

that smolts be able to reach marine waters as quickly and easily as

possible, with their passage unhindered.

1.3 | The restoration project

Barriers cause the upstream portion of the river to become inundated

and thus can hinder the passage of smolts heading downstream due to

FIGURE 1 Map of River Villestrup. (a) River Villestrup is situated in northentering the Mariager Fjord. (c) Seven weirs were present in the system orican be viewed at wileyonlinelibrary.com] 99

the slowing of water (e.g., Schwinn, Aarestrup, Baktoft, & Koed, 2017)

and difficulties associated with finding a safe passage route past the

structure itself (e.g., Thorstad, Økland, Kroglund, & Jepsen, 2003). Fur-

thermore, barriers hinder the upstream passage of adult trout during

their spawning migration. In Denmark, such barriers often occur in

the form of weirs in conjunction with fish farms. River Villestrup

(northeast Jutland, Denmark) historically had 17 fish farms. In an

attempt to restore the river to its original state and reinstate connec-

tivity on the lower two thirds of the river, six weirs (five in the

mainstem and one in a tributary) were removed. All associated fish

farms were simultaneously closed. The weirs were likely to have been

several hundred years old, though precise years of origin are not avail-

able. Each weir was originally made of concrete or wood and removed

by digging and removing all parts of the structure completely. Each

removal occurred within the course of a few days, though weirs were

removed in different years. In 2004, when the restoration project

began, seven weirs were left. The lowermost weir was removed in

2005, and five more were subsequently removed between 2010 and

2013 (see Figure 1c for weir locations and Table 1 for specific details

on each weir). Today, only one weir remains in the upstream portion

of the river (Figure 1c, no. 6). This study investigated the effectiveness

of this restoration approach with regard to the smolt run over the

course of 12 years (five study years).

2 | MATERIALS AND METHODS

2.1 | Study site and trap set up

River Villestrup is located in northeast Jutland (Denmark), where it

runs for 20 km before entering the Mariager Fjord (Figure 1). The river

is fed by groundwater and rainfall and has a mean annual discharge of

1.1 m3 s−1. It is home to a wild population of partially anadromous

brown trout, with both resident and migratory phenotypes. Before

the weir removals, river Villestrup was characterized mostly by sandy

and muddy substrates in the close vicinity of the weirs, with little

pool/riffle habitat. As in several Danish rivers, river Villestrup had

and still has a relatively low gradient (approximately 1.0%) and

meandering form. However, following the removals, the river bed is

‐eastern Jutland, Denmark. (b) It runs for approximately 20 km beforeginally; six were removed, with one still remaining (no. 6) [Colour figure

TABLE 1 Weirs in River Villestrup

Weirno.

Height(m)

Width(m)

Length of pondedzone (m)

Fishwaypresent?

Year ofremoval

1 1.9 5.9 800 Yes 2005

2 1.8 4.1 180 Yes 2012

3 0.1 5.6 0 Yes 2012

4 1.8 2.7 600 Yes 2012

5 1.5 5.0 600 Yes 2012

6 1.8 4.8 900 No Not removed

7 1.0 1.7 500 Yes 2012

Note. Height (m), width (m), length of ponded zones (m), presence orabsence of fishway, and date of removal for the weirs found in RiverVillestrup.

BIRNIE‐GAUVIN ET AL. 3

characterized by coarse, gravelly substrates. For every study year (i.e.,

2004, 2008, 2009, 2015, and 2016), a full‐covering Wolf trap (8‐mm

grid spacing; Wolf, 1951) was set up 200 m from the mouth of the

river (Figure 1c, no. 1). The trap covered the entire width of the river

(approximately 6 m), allowing us to capture virtually every down-

stream migrating fish larger than 10 cm. The trap was in place from

April 1 to May 31 every year and was emptied daily during that period.

Unfortunately, given the expenses and time required to maintain a

trap for 2 months, we could not perform the study continuously

between 2004 and 2016. Thus, specific study years were selected to

provide the most representative data to evaluate the effects of weir

removal through a before‐after approach.

2.2 | Fish processing

Every day during the study period, the trap was emptied to count and

measure (±0.1 cm) all smolts. Fish were anaesthetised with benzocaine

(0.03 g l−1) for measurements and fin clipped (adipose fin). Fish were

then released just downstream of the trap. Although it was unlikely, fish

could return upstream after having been measured. In that case, fin‐

clipping allowed us to detect if a fish had already been measured and

counted, and that individual was then removed from the day's count.

2.3 | Environmental variables

Water discharge data were obtained from a monitoring station located

750 m upstream of the trap. Temperature data were obtained using an

underwater temperature data logger (Onset HOBO Tidbit v2 UTBI‐

001, range: −20°C to 70°C, Massachusetts, USA).

TABLE 2 Smolt output

2004 2008

Average length (cm) 16.3 ± 3.0 15.5 ± 4.2

Average daily count 27.2 75.4

Total count 1660 4598

Most in a day 92 931

Note. Average length of brown trout (Salmo trutta), average daily count, total c

2.4 | Data analysis

All trout between 10.0 and 21.0 cm caught in the trap were consid-

ered to be smolts (despite coloration) for the purpose of the analysis.

This is a fair assumption given the close distance between the trap and

the fjord. Furthermore, a follow‐up electrofishing pass downstream of

the trap after the end of the smolt season showed very few trout.

Mean length between years was compared using a simple linear

regression model: log lenghtið Þ ¼ yeari þ εi;

εi∼N 0; σ2� �

:(1)

Lengths were log‐transformed to meet assumptions of normality

and homoscedacity.

All statistical analyses were performed using R version 3.4.1.

3 | RESULTS

The size of the smolt run increased following the removal of weirs,

with the largest class in 2015, followed by 2016 (Table 2; Figure 2).

Average length of downstream migrating trout was different across

study years, decreasing significantly every year (p < .05; Figure 3).

We note an indication of earlier peak migration following weir removal

(Figures 2 and 4).

4 | DISCUSSION

4.1 | Smolt run

The removal of low‐head weirs in River Villestrup strongly increased

smolt output. The removal of the most downstream weir in 2005

alone led to a large increase in smolts in 2008 and 2009, suggesting

that reestablishing the ease of access to the fjord aided a large number

of fish in successfully migrating to marine environments. Given that a

Danish smolt cohort typically resides in freshwater for 1 to 2 years

before migrating, the timeline of these observations are in line with

the prediction that the effects of weir removal may take 2+ years to

appear, though we do not have data for the years of 2006–2007 to

demonstrate this. The subsequent removal of five more weirs led to

an even greater increase in 2015 and 2016. Our results indicate that

weir removal reinstated the natural habitat of the river, with many

areas dominated by fast‐moving water, riffles, and coarse substrate,

where ponded zones previously were. These environmental changes

presumably restored or even created new grounds ideal for spawning

and early development which adults and fry did not have access to for

centuries, when fish farms and mills were first established in the river

system. Adult sea trout are also able to spawn farther upstream than

2009 2015 2016

14.5 ± 23.6 13.3 ± 2.4 13.2 ± 2.2

82.6 312.9 134.2

5038 19105 8185

263 5214 1853

ount and most caught in a single day for each study year.

FIGURE 2 Catch per day. Number of downstream migrating brown trout (Salmo trutta) smolts and discharge (m3 s−1) in River Villestrup betweenApril 1 and May 31, for years (a) 2004, (b) 2008, (c) 2009, (d) 2015, and (e) 2016.

4 BIRNIE‐GAUVIN ET AL.

01

when the barriers were present. Preliminary data shows a 9‐fold

increase in adult spawners between 2004 and 2016 (from an esti-

mated 333 to 3700 individuals, data unpublished). Furthermore,

observations also suggest that sedimentation caused by barriers may

trap fry upon emergence (Rubin, 1998). The removal of obstacles

would then also increase the survival of fry, and thus, result in a larger

smolt run. Unfortunately, our set up did not allow us to follow sedi-

ment displacement post‐removal, and we cannot exclude the possibil-

ity of sediments being deposited on spawning grounds. However,

observations from fisheries technicians and local anglers supported

an increase in the number of spawning grounds throughout the river

length, with a large increase in sea trout spawners. We therefore sug-

gest that the increase in availability of spawning grounds may have

offset the negative impacts of sediment release caused by the

removals. Observations of increased spawners suggest that even if

sediments ended up on spawning grounds, the effects were

nonproblematic. 1

We observed an unexpected decrease in the smolt output

between 2015 and 2016. Three possible explanations arise. (1) It is

possible that the large smolt run from 2015 reduced the smolt output

from 2016. Previous research has shown that the density of an age

class of brown trout can affect one or more subsequent age classes

through intraspecific competition between cohorts (Elliott, 1994;

Nordwall, Näslund, & Degerman, 2001). In this case, the 1+ age class

which migrated in 2015 may have significantly reduced the abundance

of the 0+ age class which would have migrated in 2016, either through

predation, density‐dependent mortality, or intraspecific competition

(Elliott, 1994). (2) It is possible that the decrease was due to variation

in the annual smolt production, which may vary from year to year due

to variation in biotic and abiotic factors (Chadwick, 1982; Warren,

Dunbar, & Smith, 2015). In this case, we would expect the number

of smolts to increase again in the upcoming years. (3) It is possible that

the population suffered high overwinter mortality due to harsh envi-

ronmental conditions (Elliott, 1993). At least one other Danish stream

FIGURE 3 Length distribution. Left: Lengthdistributions of downstream migrating browntrout (Salmo trutta). Red dots and intervalsindicate mean length ± SD. Right: Visualizationof the fitted model. Estimated mean lengthand associated 95% confidence intervals(back‐transformed to original scale). Mean log(length) were significantly different betweenall years (p < .05) [Colour figure can be viewedat wileyonlinelibrary.com]

FIGURE 4 Migration timing. Cumulative migration curve for browntrout (Salmo trutta) smolts for each study year [Colour figure can beviewed at wileyonlinelibrary.com]

BIRNIE‐GAUVIN ET AL. 5

102

was found to have poor overwinter survival (personal observation, K.

Birnie‐Gauvin, Gudsø stream).

4.2 | Smolt size and peak migration

We observed a decrease in the average smolt size through the years.

It is possible that following weir removal, smaller fish were also suc-

cessful in migrating downstream, rather than larger fish only, which

are presumably more apt at escaping predators in ponded zones or

overcoming weirs (Winstone, Gee, & Varallo, 1985). In other words,

smaller fish no longer get stuck at weirs and/or penetrate the grid

used to prevent fish from entering the water intake channel at fish

farms and are capable of descending downstream. Another possibil-

ity for progressively smaller fish following weir removal is that a

greater number of fish caused higher intraspecific competition for

food and may have resulted in smaller fish (Holm, Refstie, & Bø,

1990). Additionally, it is likely that spawning success and recruitment

increased, which simply increased the number of migrating fish, with

a wide range of sizes. Our findings likely reflect a combination of all

three possible explanations. Alternatively, it is possible that the

removals impacted the invertebrate community, and thus, may have

reduced food availability. Although we cannot rule out this explana-

tion, it is rather unlikely that the post‐removal invertebrate commu-

nity had diminished so much that fish were smaller. Because fast‐

flowing water is typically inhabited by different invertebrate types

than slower moving water (Doisy & Rabeni, 2001), we argue that

6 BIRNIE‐GAUVIN ET AL.

the invertebrate community changed rather than diminished post‐

removal.

We expected the peak migration to occur earlier following the

removal of the weirs through a reduction in delays at ponded zones

but cannot make that conclusion for certain. Although our results indi-

cate a trend for an earlier peak migration, flood events during the

study years make it impossible to make a meaningful analysis. Evi-

dence suggests that dams delay the passage of migrating fish greatly

(Aarestrup & Koed, 2003; Gauld, Campbell, & Lucas, 2013), and that

these effects are worse when multiple dams must be overcome

(Caudill et al., 2007). Ponded zones can cause smolts to lose their ori-

entation due to diminished flow, thus delaying them (Schilt, 2007). The

removal of five of the six weirs in the main stem of river Villestrup

likely prevented such delays in downstream migration, thus enabling

fish to reach marine environments faster.

03

4.3 | Implications

Our results suggest that complete barrier removal has several impor-

tant implications for freshwater fisheries and river management. Weir

removal presumably increases the number of adult fish able to suc-

cessfully migrate upstream and spawn, perhaps due to a reduced inci-

dence of injuries at obstacles, diminished energy expenditure to attain

spawning grounds (i.e., adults no longer have to invest energy to sur-

pass barriers), and by making impassable stretches into passable ones

(Castro‐Santos & Letcher, 2010). Furthermore, weir removal may

increase reproductive output through successful egg emergence (i.e.,

unhindered by sedimentation), which would then lead to an increased

recruitment rate and an increased smolt output in the following 2+

years. Weir removal also makes smolts more successful in their down-

stream migration via reduced mortality at fish farm intake grids

(Aarestrup & Koed, 2003), reduced predation at ponded zones

(Jepsen, Aarestrup, Økland, & Rasmussen, 1998), decreased delays

(Aarestrup & Koed, 2003; Schilt, 2007), and presumably decreased

energy expenditure. In addition, barriers may induce an artificial popu-

lation structure by favoring larger individuals; removal can reinstate a

more natural population structure, with a wider size range.

Many of the fish species that migrate between freshwater and

marine waters, including trout, are used as indicator species for good

environmental and ecological status, as they experience many habitats

during their movement from upland streams to lowland rivers and

then to the sea (Gough, Philipsen, Schollema, & Wanningen, 2012;

Lasne, Bergerot, Lek, & Laffaille, 2007). Their importance in the con-

text of management cannot be understated. Fish usually migrate for

one of three reasons; migrations are either for spawning, feeding, or

refuge seeking (Northcote, 1984). Regardless of the causes for migra-

tion, barriers diminish the ease of access to spawning and feeding

grounds and hinder passage to refuge areas. These effects are likely

exacerbated in rivers with numerous barriers (Lucas & Batley, 1996).

Extensive fragmentation of river connectivity limits dispersal of many

fish species (McLaughlin et al., 2006). Furthermore, dams impact the

hydrogeomorphology of streams in some places. For example, barriers

cause a decrease in water velocity, an increase in water temperature, a

decrease in oxygen availability, and sedimentation (Baxter, 1977;

Petts, 1984). Because most diadromous species exhibit homing 1

behaviour, and because the latter is directly related to predictable

environmental conditions such as temperature, water chemistry, and

rhythmic patterns of environmental changes, their homing behavior

is likely to be greatly impacted by the presence of obstacles (Lucas &

Baras, 2008).

In the present study, we demonstrate that weir removal is an

appropriate approach to reinstate river connectivity and to increase

long‐term population sustainability of fish species. We provide some

of the first data evaluating the full river system effects of barrier

removal and further emphasize the need to implement this approach

in management schemes whenever possible.

ACKNOWLEDGEMENTS

We wish to thank all technicians and volunteers for their help in the

field. We also wish to thank the Danish Rod and Net Fish License

Funds for providing funding to support this study. K. B. G. is funded

through the AMBER project (EU #689682).

ORCID

K. Birnie‐Gauvin http://orcid.org/0000-0001-9242-0560

H. Baktoft http://orcid.org/0000-0002-3644-4960

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How to cite this article: Birnie‐Gauvin K, Candee MM,

Baktoft H, Larsen MH, Koed A, Aarestrup K. River connectivity

reestablished: Effects and implications of six weir removals on

brown trout smolt migration. River Res Applic. 2018;1–7.

https://doi.org/10.1002/rra.3271

Manuscript VI Moving beyond fitting fish into equations: Progressing the fish passage

debate in the Anthropocene

Published in: Aquatic Conservation: Marine and Freshwater Ecosystems doi: 10.1002/aqc.2946

105

Received: 20 December 2017 Revised: 30 March 2018 Accepted: 5 May 2018

DOI: 10.1002/aqc.2946

S P E C I A L I S S U E A R T I C L E

Moving beyond fitting fish into equations: Progressing the fishpassage debate in the Anthropocene

Kim Birnie‐Gauvin1 | Paul Franklin2 | Martin Wilkes3 | Kim Aarestrup1

1Section for Freshwater Fisheries and

Ecology, National Institute of Aquatic

Resources, Technical University of Denmark,

Silkeborg, Denmark

2National Institute of Water and Atmospheric

Research, Hamilton, New Zealand

3Centre for Agroecology, Water and

Resilience, Coventry University, Ryton‐on‐Dunsmore, UK

Correspondence

K. Birnie‐Gauvin, Section for Freshwater

Fisheries and Ecology, National Institute of

Aquatic Resources, Technical University of

Denmark, Vejlsøvej 39, 8600 Silkeborg,

Denmark.

Email: kbir@aqua.dtu.dk

Funding information

AMBER (European Union), Grant/Award

Number: 689682; Danish Fishing License

Funds; KEEPFISH (European Commission),

Grant/Award Number: RISE‐2015‐690857‐KEEPFISH; New Zealand Ministry for Busi-

ness, Innovation and Employment contract,

Grant/Award Number: C01X1615

Aquatic Conserv: Mar Freshw Ecosyst. 2018;1–11.

Abstract

1. Realization of the importance of fish passage for migratory species has led to the

development of innovative and creative solutions (‘fishways’) to mitigate the

effects of artificial barriers in freshwater systems in the last few decades.

2. In many instances, however, the first move has been to attempt to engineer a solu-

tion to the problem, thus attempting to ‘fit fish into an equation’. These fishways

are often derived from designs targeting salmonids in the Northern Hemisphere.

They are rarely adequate, even for these strong‐swimming fish, and certainly appear

to be unsuitable for most other species, not least for those of tropical regions.

3. Fishway design criteria do not adequately account for natural variation among indi-

viduals, populations and species. Moreover, engineered solutions cannot reinstate

the natural habitat and geomorphological properties of the river, objectives that

have been largely ignored.

4. This article discusses the most prominent issues with the current management and

conservation of freshwater ecosystems as it pertains to fish passage. It is not

intended as a review on fish passage, but rather a perspective on the issues related

to fishways, as seen by practitioners.

KEYWORDS

biodiversity, conservation, dams, ecological engineering, fishways, freshwater, habitat, hydropower,

management, weirs

1 | INTRODUCTION

Fragmentation of freshwater ecosystems has been identified as one of

numerous global river syndromes characteristic of the Anthropocene

(Meybeck, 2003). Continued human population growth will only serve

to increase pressures on water resources, driving further investment in

infrastructure to support water, food and energy security, and to pro-

tect land and property from flooding (Garcia‐Moreno et al., 2014;

Vörösmarty et al., 2010). For example, at least 3700 major hydro-

power dams (capacity >1 MW) are planned or under construction

worldwide, and the number of smaller dams (<1 MW) planned is likely

to significantly exceed this (Zarfl, Lumsdon, Berlekamp, Tydecks, &

Tockner, 2015).

wileyonlinelibrar

Whereas ensuring access to food, energy and potable water is

fundamental for supporting the future of human societies, freshwater

biodiversity in the Anthropocene is under great threat owing to unsus-

tainable river basin development (Garcia‐Moreno et al., 2014; Poff,

2014; Vörösmarty et al., 2010). Continuing river fragmentation and

dam construction present one of the greatest global threats to fresh-

water biodiversity and ecosystem functioning (Dudgeon et al., 2006).

Disruptions to river connectivity threaten ecosystem structure and

function by interrupting movements of migratory species (Winemiller

et al., 2016), blocking the exchange of individuals and genetic informa-

tion between populations (Raeymaekers et al., 2008; Wofford,

Gresswell, & Banks, 2005), modifying aquatic habitats and altering

flow and sediment transport regimes (Bunn & Arthington, 2002).

© 2018 John Wiley & Sons, Ltd.y.com/journal/aqc 1

2 BIRNIE‐GAUVIN ET AL.

Unfortunately, consideration of biodiversity and ecosystem function-

ing tends to take a distant second place to engineering solutions that

meet immediate human needs (Garcia‐Moreno et al., 2014). This is

despite the increasing recognition that biodiversity loss impairs and

fundamentally alters the functioning of ecosystems upon which soci-

ety depends for food, energy and water security (Vignieri, 2014).

Globally, freshwater fish are a critical food resource and support

economically and culturally important fisheries (e.g. Winemiller et al.,

2016). As a result, the loss of fish populations during the

Anthropocene has probably received greater global attention than

any other freshwater group. Connectivity is fundamental to the

structure and functioning of freshwater fish communities and aquatic

ecosystems worldwide, and is active along the longitudinal, vertical,

lateral and temporal dimensions (Tockner, Schiemer, & Ward, 1998).

Instream structures, such as dams, weirs, tide gates and culverts, inter-

rupt connectivity in all dimensions, with the repercussions being

observed as species or population declines and extirpations in river

systems across the globe (Table 1).

The impact of instream structures on the movements and migra-

tion of fish has long been recognized. In northern Europe, fishways

were already being established by the mid‐18th century. Although

these early fishways were inefficient (Francis, 1870), their presence

indicates the recognition of connectivity issues. At that time, the main

concern was the upstream passage of Atlantic salmon, Salmo salar,

mostly because of its high economic and recreational value (Katopodis

&Williams, 2012). Despite the ever‐increasing awareness of impacts of

barriers on other fish species (Branco, Amaral, Ferreira, & Santos, 2017;

Perkin et al., 2015; Raeymaekers et al., 2008; Wilkes, McKenzie, &

Webb, 2018), contemporary approaches to fish passage research and

management continue to be dominated by salmonid‐centric methods,

solutions and thinking, and continue to focus on the upstream passage

of fish at larger structures, giving relatively little attention to equally

important downstream movements and small structures.

Increasing realization of the importance of effective fish passage

for sustaining migratory species has led to the development over

recent decades of innovative and creative solutions to mitigate the

effects of artificial barriers in freshwater systems, but management

of fish passage continues to be dominated by an ‘impair‐then‐repair’

approach (Vörösmarty, Pahl‐Wostl, Bunn, & Lawford, 2013). For most

TABLE 1 Examples of fish population declines and local extinctions ascr

Species Location Fragme

Atlantic salmon, Salmo salar Rhine, Seine and Garonnebasins, France

Disappe

Gudenaa River, Denmark

Pacific salmon, Oncorhynchus spp. Pacific Coast, USA 101 stoof ex

Whitespotted char, Salvelinus leucomaeni Hokkaido, Japan Local eupstr

Dabry's sturgeon, Acipenserdabryanus

Yangtze River Criticall(poss

Spotted sorubim, Pseudoplatystomacoruscans

São Paulo state, Brazil Rapid loconst

Jullien's golden carp, Probarbusjullieni

Northern Malaysia Possiblyand s(Pera

dams and other instream infrastructure, fishways continue to be con-

sidered an add‐on ‘fix’ once the standard structural design is complete

(Katopodis & Williams, 2012). Furthermore, fish passage tends to be

treated on a site‐by‐site basis, focused only on getting fish from one

side of the structure to the other, and monitoring their effectiveness

is often absent. Rarely is consideration given to the broader catchment

context of fish passage, or the impacts on aquatic habitats and ecosys-

tem processes (Kemp, 2016; McLaughlin et al., 2013; Pelicice &

Agostinho, 2008; Pompeu, Agostinho, & Pelicice, 2012; Silva et al.,

2018). We argue that this reductionist approach is symptomatic of

the origins of fish passage research, embedded in a philosophy of

engineering a way out of the problems created by human modifica-

tions of the riverscape.

A characteristic of the dominant engineering approach to fish pas-

sage is determinism (e.g. ‘the species can swim at x velocity for t time’).

A general failure to consider the bigger picture and a continued focus

on trying to ‘fit fish into equations’ cannot account for the natural

variation among individuals, populations and species that is an

essential characteristic of sustainable aquatic ecosystems. To improve

outcomes for freshwater biodiversity, we believe that fish passage

research and its applications must embrace this natural variability. To

achieve this there is a need to confront what we view as inherent

biases in fish passage research, policy and practice that are derived

from the overwhelming dominance of research on the salmonid spe-

cies of the temperate Northern Hemisphere. The field of fish passage

as a whole needs rethinking, with the objective of helping fish move

up and down rivers with no adverse effects.

The aim of this article, therefore, is to contribute to the continuing

debate on fish passage (e.g. Bunt, Castro‐Santos, & Haro, 2016;

Kemp, 2016; Silva et al., 2018; Williams & Katopodis, 2016) by provid-

ing a perspective on what we view to be among the most crucial

issues related to the prevailing paradigm of fish passage research

and management at a global scale. In particular, we consider the

question of whether the current approach to the fish passage problem

is fit for purpose and suitable for effectively tackling the freshwater

biodiversity crisis of the Anthropocene. The article finishes by propos-

ing some potential approaches to progress the fish passage debate by

moving beyond some of the biases we identify, and pursuing a more

holistic approach to fish passage research and applications.

ibed to river fragmentation

ntation impacts References

arance of whole stocks Porcher and Travade (1992)

Jepsen et al. (1998)

cks at high risktinction

Nehlsen, Williams, and Lichatowich(1991)

xtinction at 17 siteseam of dams

Morita and Yamamoto (2002)

y endangeredibly extinct)

Wei, He, Yang, Zhang, and Li (2004),Wei et al. (1997), Wan, Fan, andLi (2003)

cal extinction after damruction

Welcomme (1985)

local extinction (Pahang River)ignificant population declinek River)

Baird (2006), Dudgeon et al. (2006)

FIGURE 1 Burst swimming speeds (the maximum swimming velocitythat a fish is capable of sustaining for up to 20 s) of salmonids andother migratory fish with characteristic body lengths at the time ofupstream migration. All species listed are defined as diadromous or

potamodromous in FishBase (Froese & Pauly, 2016). All studies listedsampled burst swimming speeds in laboratory flumes. Symbols showmodes. Whiskers show range from selected studies to demonstratepopulation‐level variation. Examples cited: 1Nikora, Aberle, Biggs,Jowett, and Sykes (2003); 2Plew et al. (2007); 3Rodgers et al. (2014);4Starrs, Ebner, Lintermans, and Fulton (2011); 5Colavecchia,Katopodis, Goosney, Scruton, and McKinley (1998); and 6Tudorache,Viaene, Blust, Vereecken, and De Boeck (2008)

BIRNIE‐GAUVIN ET AL. 3

08

2 | BIASES IN FISH PASSAGE RESEARCHAND APPLICATION

2.1 | Long‐standing focus on salmonids andupstream passage

Much of the knowledge we have about the effects of instream

barriers, fishways and the ability of fish to pass them is derived from

studies based on anadromous salmonids in the temperate Northern

Hemisphere. This focus emerged due to the well‐documented declines

in salmonid stocks in river systems around the globe arising from

anthropogenic interruptions to migration routes (e.g. Yeakley, Maas‐

Hebner, & Hughes, 2014). Owing to the economic and cultural

importance of salmonid populations, and often supported by local

legislative requirements, efforts to ‘fix’ the problem emerged. Despite

these efforts, there remains a focus on upstream movements, with less

consideration given to getting fish back downstream (although efforts

to address downstream movement have risen in recent years; e.g.

Arnekleiv, Kraabøl, & Museth, 2007; Birnie‐Gauvin et al., in press).

Adult salmonids have very particular needs given their highly

directed and relatively synchronized migration. Salmonid migratory

behaviour is some of the most studied, although downstream move-

ments have received considerably less attention. The behaviour of

downstream migrating salmonid smolts is often simplified, and

believed to be addressed by designing screens and bypasses that

screen fish only near the water surface (Arnekleiv et al., 2007). In

our experience, however, a significant proportion of smolts move

below the screen, with evidence of individuals migrating near the

bottom (Svendsen, Eskesen, Aarestrup, Koed, & Jordan, 2007). A lack

of focus (and knowledge) on this downstream movement, combined

with the observation of highly synchronous upstream migrations, has

led to the perception that these fish have relatively narrow and well‐

defined needs, with characteristics that suit the reductionist approach

of the engineering discipline.

Historically, designing effective fish passage solutions was chal-

lenged by the constraints (primarily space, cost and flow) typically

imposed by having to append fishways to existing structures retrospec-

tively. Solutions inevitably became a balancing act between overcoming

the fall height created by the obstruction, minimizing fishway length

and maintaining hydraulic conditions in the fishway within the capabil-

ities of the target species and life stage, and only generating marginal

changes to the function of the obstacle in question. Adult salmonids

are agile and highly capable swimmers because they swim upstream

and thus have a greater ability to overcome more hydraulically chal-

lenging environments than many other species do. This has had a

strong influence on the type and hydraulic performance standards of

most fishway designs that exist today (Mallen‐Cooper & Brand, 2007).

Fish passage research remains largely entrenched in the early para-

digm of salmonid biology. This long‐standing focus has resulted in the

same approach being perpetuated all over the globe, for all species, in

all geographical contexts, rather than taking a step back and rethinking

whether it is the right approach in a particular location (Link & Habit,

2015; Mallen‐Cooper & Brand, 2007; Wilkes, Baumgartner et al.,

2018). Despite the significant differences between the requirements

of salmonids and most other fishes (Figure 1), including those from the 1

tropics and temperate Southern Hemisphere, the knowledge, tech-

niques, thinking and solutions developed from studies of salmonids have

been widely transferred to fish passage design and management else-

where (Silva et al., 2018). Application of these approaches to freshwater

systems with native species that have completely different needs has

contributed to repeated failures and poor performance of fishways

around the world (Lira et al., 2017; Wilkes, McKenzie et al., 2018). For

example, Mallen‐Cooper and Brand (2007) showed very poor passage

of native Australian fish species through a salmonid fishway on theMur-

ray River, with <1% of themost abundant species ascending. The contin-

ued underwhelming performance of many salmonid fishways (Brown

et al., 2013), and continuing unsuccessful application of salmonid‐centric

solutions to non‐salmonid species has led some to suggest that, in a

global sense, fishways are a technology in decline (Kemp, 2016).

2.2 | Engineering a way out of the problem

The fundamental dichotomy of the fish passage problem is the need to

balance the trade‐offs between doing what would be best ecologically

(i.e. remove all barriers) and trying to engineer a way out of the problem

where there is a need for essential infrastructure (Nieminen,

Hyytiäinen, & Lindroos, 2017). In too many instances, engineered solu-

tions continue to be the default first step to solving fish passage issues.

We suggest that this bias has emerged from the emphasis of early fish

passage research on retrospectively engineering site‐scale solutions to

fix problems for individual species at existing infrastructure. This has

embedded the idea of fish passage solutions as an ‘add on’ to structural

designs, rather than an integral component of the design to be consid-

ered from the outset. However, inappropriate transfer of technological

solutions and increasing evidence of the unintended consequences of

4 BIRNIE‐GAUVIN ET AL.

providing fish passage (McLaughlin et al., 2013; Pelicice & Agostinho,

2008; Pelicice, Pompeu, & Agostinho, 2015), along with the broader

ecosystem changes (Birnie‐Gauvin, Aarestrup, Riis, Jepsen, & Koed,

2017), raise questions over the continued suitability of this approach.

Obviously, there are instances where instream infrastructure is

necessary, and hence there will always be cases where engineered

solutions are required. However, current design philosophies tend to

force ecologists to take a reductionist approach, trying to fit fish into

equations suitable for engineers to work out a solution that fits the

appropriate hydraulic design envelope and minimizes costs. This

approach has undoubtedly contributed to the less than satisfactory

success of many fish passage solutions, as evidenced in multiple

reviews (Bunt, Castro‐Santos, & Haro, 2012; Lira et al., 2017; Noonan,

Grant, & Jackson, 2012; Roscoe & Hinch, 2010). The simplified repre-

sentations of reality required by this approach, while convenient, inev-

itably fail to capture the natural variation that is characteristic of all

organisms, ecological communities and ecosystems. Furthermore, it

is impractical to effectively characterize the full range of hydraulic

requirements of multiple species and life stages of fish in sufficient

detail to provide effective hydraulic design criteria, particularly when

considering ‘megadiverse’ fish communities such as those typical of

tropical regions (Winemiller et al., 2016).

We encourage a more holistic approach, planning infrastructure

and designing structures from the outset with a view to maintaining

ecosystem processes and functioning, including aiming for the

seamless movement of organisms. Doing so requires a change in

design philosophy and a shift in expectations of how things should

be done at every level. Scientists, engineers and managers must realize

that the difference between removing (or not installing) a barrier and

constructing a fishway is huge; fishways will never be as effective as

the complete absence of barriers for providing fish with sufficient

habitat and allowing safe movement. We argue that the first question

that should always be asked (perhaps twice) is whether that barrier is

necessary at all, and if so whether a fishway will contribute to the

maintenance of viable populations upstream and downstream of

the structure (Pompeu et al., 2012). There is strong evidence that

removing artificial barriers to migration can be cost‐effective and

result in the rapid recovery of freshwater biodiversity and ecosystem

processes, as seen for American eel (Anguilla rostrata; Hitt, Eyler, &

Wofford, 2012), sea lamprey (Petromyzon marinus; Hogg, Coghlan, &

Zydlewski, 2013) and brown trout (Salmo trutta; Birnie‐Gauvin, Larsen,

Nielsen, & Aarestrup, 2017; Birnie‐Gauvin et al., in press), as well as

other species (O'Connor, Duda, & Grant, 2015), yet barrier removal

remains relatively uncommon, even where structures are redundant.

Consequently, despite the growing use of fishways, which are suppos-

edly designed to allow migrating fish to bypass barriers and reach suit-

able habitat in which to grow and reproduce, these structures remain

mere pacifiers of the underlying ecological problems (Bunt et al., 2012,

2016; Lira et al., 2017; Noonan et al., 2012; Roscoe & Hinch, 2010).

09

2.3 | Requirement mismatches and ignoring naturalvariation

The dominance of salmonid studies and reductionist approaches to

engineering design have combined to result in a situation where 1

considerations of natural variations in fish behaviour and dispersal

capabilities are minimized. Migration is a concept that has been known

and studied for centuries. Its occurrence is widespread across all major

taxonomic groups and has piqued the interest and curiosity of scien-

tists for as long as it has been known. For decades, fish biologists have

tried to understand its underpinning mechanisms and drivers, making a

point of protecting migratory species as they usually depend on at

least two types of environments to thrive (e.g. eels growing in fresh

water and migrating to salt water to spawn). Although many of the

overarching concepts of migration are well known, and largely

accepted, the focus on a relatively narrow range of high‐status species

has biased management actions towards particular life‐history strate-

gies. Furthermore, it has led to a halt in questioning some of the basic

information regarding migration.

The majority of fish passage solutions have been designed to

cater for anadromous life histories. Even within the well‐studied

salmonid species, however, there is growing evidence that salmonid

smolt migrations occur throughout the year rather than during a single

peak period (Aarestrup, Birnie‐Gauvin, & Larsen, 2018; Winter,

Tummers, Aarestrup, Baktoft, & Lucas, 2016). Despite this, current

strategies for fish passage management, such as spillway opening

and dam or weir closure periods, typically only occur during the peak

spring migration for smolts, neglecting to cater for fish that do not

fit the currently accepted salmonid paradigm (Aarestrup et al., 2018).

Another important consideration is the ‘migratory’ versus ‘non‐

migratory’ or ‘resident’ terminology; it creates the perception that

non‐migratory or resident fish do not move. Yet they do (Jepsen &

Berg, 2002; Radinger & Wolter, 2014; Schlosser & Angermeier,

1995), and they may be affected by barriers more than is traditionally

recognized (Branco et al., 2017). The whole fish passage issue has

largely focused on obligate migrants, sometimes classifying facultative

migratory species as non‐migratory for the purpose of passage needs.

The functional explanations for movement of ‘non‐migratory’ or ‘resi-

dent’ fish are manifold and may involve distances of the same order of

magnitude to those characteristic of ‘migratory’ species. The reasons

include: (i) to avoid unpredictable resource scarcity and perturbances

(Falke, Fausch, Bestgen, & Bailey, 2010); (ii) to repopulate habitats

previously affected by disturbance or disease (Perkin et al., 2015);

(iii) to shift distribution gradually in response to large‐scale environ-

mental change, including climate change (Hari, Livingstone, Siber,

Burkhardt‐Holm, & Guttinger, 2006); and (iv) to exchange adaptive

genetic information in the face of environmental change (Brauer,

Hammer, & Beheregaray, 2016). We stipulate unpredictability in

some of the instances listed herein because if the phenomena were

predictable the species may well be considered migratory. Such

‘unpredictability’ also encompasses the effects of climate change, so

movement for resident fish is likely to become even more important.

There is a need in the first instance, therefore, to recognize this diver-

sity of movements that occur within and between species and over

time, and to cater for this in fish passage research and applications.

There is also a need to consider variation at the individual level.

Individuals vary in their ability and motivation to overcome

barriers (Agostinho, Pereira, Oliveira, Freitas, & Marques, 2007; Bunt

et al., 2012). There is also variation among populations of the same

species (Birnie‐Gauvin, Larsen, Thomassen, & Aarestrup, 2018;

BIRNIE‐GAUVIN ET AL. 5

10

Figure 1). The reductionist approach typically adopted for fishway

design means that this natural variation is often neglected completely,

or is at least poorly accounted for (but see Wilkes, Baumgartner et al.,

in press). Variation in fish behaviour and requirements is wide ranging,

and often discounted in modelling exercises, potentially rendering the

outcomes invalid when they are applied to real‐life situations.

Although modelling is a valuable tool, explicit considerations of the

uncertainty created by natural variation need to be implemented.

Most modelling approaches in fish passage research, at their core,

are equations. This means that fish must be fitted into a mathematical

phrase, essentially collapsing all natural variation into one ‘magic’

number, even in situations where swimming behaviour between pop-

ulations is strongly divergent (Link et al., 2017). Whereas the biologist

would be calling for explicit recognition of this divergent swimming

behaviour in fishway design, the engineer may instead consider an

equation that does away with this variability.

The requirement to fit fish into equations in a way that is consis-

tent with typical engineering design practices has seen an emphasis on

efforts to quantify fish swimming speeds. The most convenient way of

achieving this is through controlled laboratory swimming tests. The

criteria for water velocity design of fishways are typically determined

through controlled swimming tests that force fish to swim at a fixed

mean velocity (endurance tests) or at an incrementally increasing

velocity (critical swimming tests) (Beamish, 1978). Although practical,

this raises several issues related to individual variability; for example,

turbulence and fish acceleration and deceleration are often ignored

(but see Plew et al., 2007); the difference between different measures

of swimming performance remains unclear (Peake, 2004); variations in

swimming performance at different temperatures or under varying

water quality are often not considered (but see Bannon & Ling,

2003); and species and individuals that do not ‘cooperate’ by swim-

ming in the laboratory are often selected out rather than being consid-

ered a separate behaviour class to be accounted for (Santos, Pompeu,

& Martinez, 2007). Furthermore, the behaviour of fish in an artificial

laboratory set‐up is unlikely to be natural because of the stress of han-

dling and the change in behaviour that comes with being held in cap-

tivity for long periods, as well as the absence of natural environmental

heterogeneity or migration cues (Vrieze, Bjerselius, & Sorensen, 2010).

This has led some workers to suggest that volitional swimming speed

tests – for example, measured in open channel flumes – are more

appropriate (Haro, Castro‐Santos, Noreika, & Odeh, 2004). However,

although this may improve the biological realism of fish swimming per-

formance evaluations, it still does not overcome the challenge of

effectively characterizing the natural variability in performance

between individuals and populations and translating them into practi-

cal design criteria that account for this uncertainty. Whereas general

relationships between hydraulics and swimming behaviour can be

investigated, and are essential for supporting development of hydrau-

lic design criteria, laboratory studies alone are insufficient for develop-

ing absolute criteria, and much greater effort should be placed on

incorporating natural variation and uncertainty into the results.

As attention in fish passage research begins to move towards

catering for multispecies assemblages, a further challenge emerges in

trying to account also for the variation between and among species

and life stages. In all but the most extreme cases, fish passage must 1

be available for more than a single species, each with potentially

different requirements, at different life stages. How can the range of

individuals that must overcome barriers be accommodated? A mature

female on her way to spawn is full of eggs. Are her swimming abilities

reduced? How can fish passage infrastructures accommodate her?

2.4 | Ignoring small‐scale barriers

The impacts of large dams have been well documented and have often

been the primary focus of fish passage research. In most river basins,

however, small‐scale structures such as weirs and culverts frequently

make up the vast majority of obstructions (Gibson, Haedrich, &

Wernerheim, 2011). Small structures, with fall heights as little as

50 mm, can be a complete barrier for some fish species (Baker,

2003), particularly the small‐bodied species characteristic of many

Southern Hemisphere fish communities (Link & Habit, 2015). Despite

their widespread distribution, these smaller barriers continue to

receive relatively little attention, as individually they are often deemed

to have small effects (Branco et al., 2017). However, there is increas-

ing evidence of their impacts on fish movements (Branco et al.,

2017; Lucas, Bubb, Jang, Ha, & Masters, 2009), and it has been sug-

gested that the cumulative effects of multiple barriers can be at least

as severe as those of large dams (Cooke et al., 2005).

Fish passage through culverts has received some attention, again

focused almost exclusively on salmonids. Early work investigated the

hydraulic effects of culvert baffling (Ead, Rajaratnam, & Katopodis,

2002; Rajaratnam, Katopodis, & Lodewyk, 1988), and more recent

studies have included observations of fish behaviour during culvert

passage (Goerig & Castro‐Santos, 2017). However, there is a need

to develop solutions appropriate to the target species. For example,

David, Tonkin, Taipeti, and Hokianga (2014) investigated a novel

approach for facilitating upstream passage of small‐bodied fish

through culverts using mussel spat ropes as baffling media, showing

that culvert passage success could be significantly improved. We sug-

gest that an increased focus on fish passage at small‐scale structures

has the potential for rapid and cost‐effective biodiversity gains. For

example, there are several studies from Australia and New Zealand

describing positive outcomes for non‐salmonid fish species richness

and abundance resulting from retrofitting fish passage solutions to

culverts (Amtstaetter, O'Connor, Borg, Stuart, & Moloney, 2017;

David & Hamer, 2012; Franklin & Bartels, 2012). Erkinaro, Erkinaro,

and Niemelä (2017) also demonstrated increases in the distribution

of juvenile Atlantic salmon following the restoration of impassable

road culverts in Finland. These approaches, however, remain embed-

ded in the philosophy of trying to engineer a fix that can be applied

to a structure rather than taking a more holistic approach to fish

passage management.

We suggest that, more importantly, small‐scale barriers also offer

the best opportunity for overcoming the bias towards engineered fish

passage fixes. Many small‐scale structures are now redundant, no

longer serving their original purpose, but are seen as valuable parts

of cultural heritage. There are obvious opportunities here for removal,

yet fish passage frequently takes a back seat to cultural interests. Very

often, the basis of local arguments that can be observed or noticed in

some way (e.g. the sound of a waterfall, a bridge over a dam, or a

6 BIRNIE‐GAUVIN ET AL.

reservoir) win over the problems that cannot be seen by the naked eye

(i.e. the fish). But what benefits are conferred from enjoying the sound

of a waterfall? Should these arguments take precedence over the pro-

tection of freshwater biodiversity? The sad reality is that, in the case

of small barriers, these arguments will often hold. Removal of such bar-

riers is often achievable and cost‐effective, and should be a priority for

achieving rapid, sustained recovery of freshwater communities

(although we acknowledge that dams can sometimes serve as a barrier

to the spread of non‐native species; Gangloff, 2013). Removal also has

the advantage of restoring physical habitat and ecosystem processes

(Birnie‐Gauvin, Aarestrup, et al., 2017; Birnie‐Gauvin, Tummers, Lucas,

& Aarestrup, 2017; Timm, Higgins, Stanovick, Kolka, & Eggert, 2017).

Under circumstances where removal is not an option, it is also

feasible and practicable to rethink design approaches to better

accommodate the unhindered movement of organisms and maintain

ecosystem processes. A good example has been the adoption of the

stream simulation approach to culvert design (Forest Service

Stream‐Simulation Working Group, 2008). This approach adopts a

more holistic method with the objective of maintaining continuity of

physical habitat and ecosystem processes between the upstream and

downstream reaches. As such, the conditions inside the culvert repli-

cate adjacent stream reaches and represent no greater impediment

to the movement of organisms than progress through the normal

stream environment. Studies of culverts built using this approach

indicate that not only do they provide effective fish passage, but

they are also more effective at maintaining sediment transport

(Timm et al., 2017), and they are more resilient to large floods than

traditional hydraulic culvert designs (Barnard, Yokers, Nagygyor, &

Quinn, 2015; Gillespie et al., 2014). It has also been shown that the

relatively modest increases in initial investment to implement stream

simulation designs can yield substantial societal and economic benefits

in the long term (Gillespie et al., 2014).

11

2.5 | More than just safe passage: Critical habitatavailability and distribution

Barriers have received so much attention largely because they hinder

the movements of fish by reducing connectivity (Wheeler,

Angermeier, & Rosenberger, 2005), and also because they alter

hydrological and thermal processes (Bergkamp, McCartney, Dugan,

McNeely, & Acreman, 2000). However, the modification and loss of

aquatic habitats caused by the presence of barriers is an impact that

is often neglected (Birnie‐Gauvin, Aarestrup, et al., 2017; Franklin &

Hodges, 2015). Although the knowledge that habitat alterations are

in fact induced by barriers is common, addressing the implications of

losing ecologically relevant habitat is rare. Because dams are most

often established in river reaches with high gradients, there can be a

disproportionate loss of rheophilic (i.e. fast‐flowing and highly‐oxy-

genated) habitat. These areas are essential for rheophilic fish species

such as salmonids and eels that depend on these ‘critical habitats’ to

complete their life cycles. Consequently, even if those species can

overcome a barrier, population viability is still compromised owing to

the loss of adequate habitat (Birnie‐Gauvin, Aarestrup, et al., 2017).

Tide gates also have a significant impact on physical habitats, reducing

hydrological exchange and interrupting natural salinity gradients, 1

in addition to blocking fish movements (Boys, Kroon, Glasby, &

Wilkinson, 2012; Franklin & Hodges, 2015). Fish survival is also

severely reduced by habitat modifications. Large predatory species,

such as the pike (Esox lucius), can thrive in impoundments, with

younger fish as a source of food (Jepsen et al., 1998). Habitat loss

should, therefore, be addressed through hydrological and morpholog-

ical mitigation, either before or simultaneously (at the very least) with

the issue of fish passage (Birnie‐Gauvin, Aarestrup, et al., 2017).

The complexity of the fish passage problem in Neotropical South

America, Southeast Asia and Africa, reflecting the diversity of native

species assemblages and the wide range of fish life‐histories there,

has highlighted the need to consider the distribution of critical habi-

tats on either side of a barrier (Pompeu et al., 2012). This broader

approach was necessary because fishways were found to be failing

as a conservation tool; high percentages of fish approaching the fish-

way were passing only to be trapped without access to critical habitats

upstream owing to reservoirs or the presence of other barriers

without fishways (Pelicice & Agostinho, 2008; Pelicice et al., 2015).

In Brazil, therefore, far from protecting fish populations, policies that

require the provision of fish passage at dams have in some cases been

the main threat to their viability (Pelicice et al., 2017).

2.6 | Lack of post‐implementation monitoring: Howwell does it work?

In many cases, monitoring the effectiveness of fishways is not imple-

mented or is not a licensing requirement. In other words, asking how

well it works is not part of fulfilling requirements, and thus post‐imple-

mentation monitoring remains unaccomplished. This is a major reason

for the unsustainable policies prevailing in Brazil, as introduced in the

previous example (Pelicice et al., 2017), and probably in many other

parts of the world. Part of the answer to this paradox relates to the

deterministic tradition of engineering, as previously discussed. If

the effectiveness of fishways is predetermined, then monitoring and

adaptive management is optional. There is rarely a statutory obligation

to prove that a fishway is achieving its overall goal of sustaining viable

fish populations, although it may be achieving other goals, such as

those associated with corporate social responsibility. Yet why have

measures in place for fish passage if there is no answer to how

many individuals get through and whether that is sufficient to sustain

fish communities?

Herein lies a critical challenge both for fish passage scientists and

practitioners: How can objectives for fishways (or more broadly for

maintaining connectivity) be defined that are ecologically meaningful

but which are also practicable (i.e. specific and measurable)? The lack

of post‐implementation monitoring is a lost opportunity. Understand-

ing how existing mitigation efforts work and do not work may yield

significant information that will help improve future rehabilitation

efforts (Birnie‐Gauvin, Tummers, et al., 2017). To achieve this, how-

ever, there needs to be guidance on what to monitor and how, and

this relies on having clearly defined objectives. Definitions such as

‘effective’ or ‘free’ fish passage can be ambiguous, open to interpreta-

tion or unachievable. The term ‘free’, for example, is frequently used to

describe fish passage targets, but this is highly unlikely to be measur-

able given the general lack of knowledge on how many fish have

BIRNIE‐GAUVIN ET AL. 7

attempted to pass a structure compared with how many fish actually

did so. Furthermore, the term ‘free’ requires that fish are not delayed,

which is seldom the case. Delay may in fact have carry‐over effects

that may lead to future adverse consequences (McCormick et al.,

2009). So, can fish passage ever be free? Yes, if the barrier is removed,

but no if a fishway is present.

Perhaps the correct scientific question to ask, therefore, is: How

many individuals that attempt to pass actually pass? Along similar

lines, the appropriate management question to ask may be: How

many individuals need to get through to meet ecological objectives

and to ensure population viability? Despite their necessity in the

context of fish passage, these questions are almost never asked,

let alone answered. Instead, there is almost invariably a focus on

the movement of individual fish in the immediate vicinity of the

structure to be passed. This focus is made possible through the use

of biotelemetry, which has emerged as the ‘gold standard’ in fish

passage research (Bunt et al., 2012; Silva et al., 2018). Use of these

techniques has undoubtedly resulted in significant advances in fish

passage science by improving understanding of the behavioural and

motivational aspects of fish movements (Aarestrup, Lucas, & Hansen,

2003). However, while continuing miniaturization of the tags used in

biotelemetry studies has broadened the size range of fish to which

this technology can been applied (Baker, Reeve, Baars, Jellyman, &

Franklin, 2017), small‐bodied fish, and fish that migrate during early

life stages (larval and juvenile), remain outside the reach of these

technologies. Consequently, if biotelemetry methods continue to be

upheld as the standard by which fish passage success is to be mea-

sured, there is a risk yet again of perpetuating the focus on larger fish

species at the expense of considering all parts of the fish community

and all life stages.

12

3 | DISCUSSION

Awareness of the impacts of instream infrastructure on fish move-

ments, and hence fish populations, has increased considerably over

the last couple of decades. Despite this, the reductionist, salmonid‐

centric, impair‐then‐repair approach to infrastructure design largely

continues to prevail, and continues to be biased towards upstream

movement. We suggest that this stems from the roots of fish passage

research emerging from attempts to retrospectively engineer fishways

as fixes for moving individual iconic species upstream at existing

infrastructure to mitigate for an emerging problem. Although we

acknowledge the significant progress that has been made in restoring

fish passage following this approach, including the benefits of studying

salmonids in this context, the effectiveness of many of these struc-

tures remains too small to be ecologically meaningful. For example,

several recent meta‐analyses have attempted to evaluate the effec-

tiveness and performance of fishways (Bunt et al., 2012; Noonan

et al., 2012; Roscoe & Hinch, 2010). The most consistent messages

that emerge from these reviews are the overwhelming dominance of

studies focusing on anadromous salmonids, and the high variability

(ranging from near 0% to near 100%) in fishway performance. As the

focus has increasingly turned to non‐salmonid fishes and catering for

multispecies assemblages in fishways, evidence of failures in the 1

current fish passage paradigm continues to mount. Largely precipi-

tated by the direct transfer of findings from the Northern Hemisphere

to diverse geographical and ecological contexts, repeated failures

and the emergence of unintended consequences has undermined

confidence and the willingness of practitioners to invest in

implementing fish passage solutions (Harris, Kingsford, Peirson, &

Baumgartner, 2016).

Although potentially disheartening, we believe that this reflects a

failure in the discipline to recognize adequately and move beyond

inherent biases in methods and ways of thinking, rather than a flaw

in the concept of fish passage itself. We are encouraged by recent

contributions to the fish passage debate, particularly emerging from

the Southern Hemisphere and the tropics, which challenge some of

these biases. Pompeu et al. (2012), for example, proposed that fishway

efficiency should be assessed based on the capability of the structure

to maintain viable fish populations, rather than a simple metric of the

proportion of fish that ascend a structure. Traditional passage effi-

ciency metrics may have been suitable for species similar to salmonids

that exhibit relatively synchronous, seasonal and highly directed

movements between clearly separated critical habitats (Kemp, 2016),

but transferring these metrics to species and populations with more

diverse life‐histories and behaviours may not be the most appropriate

measure of fish passage success. Impoundments upstream of dams

can act as ecological traps (Pelicice & Agostinho, 2008; Pelicice

et al., 2015) preventing downstream movement of eggs and larvae

necessary to complete fish life cycles. Providing effective upstream

passage for adults past dams, therefore, acts as a population sink with

adverse consequences for the long‐term sustainability of fish popula-

tions (Pelicice & Agostinho, 2008). Likewise, in New Zealand, juvenile

eels (Anguilla dieffenbachii and Anguilla australis) are regularly trans-

ferred upstream of hydropower dams to seed upstream populations,

but in most cases there is no, or only very limited, facility for subse-

quent downstream passage of migrant adults through the dams

(Jellyman, 2007). Thus, while they do support fisheries, the long‐term

value to biodiversity conservation may be questionable.

Harris et al. (2016), in a review of barrier mitigation efforts in

Australia, also highlighted the challenges of catering for a mixture of

life‐history strategies across freshwater fish communities. They pro-

posed that there is a need for river‐basin‐scale management strategies

that integrate fishway construction, where appropriate, with other

approaches, such as barrier removal, improved barrier management,

environmental flow provision and strategic prioritization of mitigation

efforts. Furthermore, they also supported the idea of broader defini-

tions of fishway success and the need for performance to be assessed

against predetermined, comprehensive biological criteria, including

considerations of the cumulative effects of multiple barriers. The

concept of river‐basin‐scale decision‐making was also emphasized by

Winemiller et al. (2016), who advocated more integrated and strategic

planning of dams that also takes into account the cumulative effects of

multiple structures on hydrology, sediment dynamics, ecosystem

productivity, fisheries and biodiversity.

We echo these calls for the need to take a step back and consider

strategies for managing connectivity at a broader scale, rather than

thinking about fish passage on a site‐by‐site basis in isolation from

the wider catchment context, as is commonly done today. Crucial to

TABLE 2 Recommendations to address biases in fish passageresearch and applications

1. Avoid building new barriers whenever possible; if unavoidable, buildthe dam/weir/culvert such that it is not a barrier.

2. First choice should always be to remove existing structures ratherthan to engineer a solution

3. Reconsider removing barrier (#2).

4. Recognize and embrace diversity of fish movement ecology.

5. Integrate natural variation and build in uncertainty to designs.

6. Use a more holistic approach, including the consideration ofgeomorphological and hydrological processes.

7. Stop recommending absolute design criteria from laboratoryswimming tests. Laboratory experiments are excellent tools forcomparative studies, but lack biological and environmental realism.

8. Use an evidence‐based approach.

8 BIRNIE‐GAUVIN ET AL.

progressing the fish passage debate is also the need to move beyond

the idea that fishways provide a universal solution to mitigating the

impacts of instream structures on aquatic communities (Brown et al.,

2013; Kemp, 2016). We do not disagree with the view of Williams,

Armstrong, Katopodis, Larinier, and Travade (2012) that with sufficient

investment in ecohydraulic research effective fishways can be

engineered, but this belief is still predicated on the anthropocentric

impair‐then‐repair approach and the assumption that providing fish

passage at instream infrastructure is inherently good. In addition, as

Kemp (2016) rightly identified, in many cases, and for the majority of

species, knowledge is currently far short of being able to develop such

technical solutions (Wilkes, McKenzie et al., 2018), and that sufficient

funding and many years of research will be required to fill those

knowledge gaps. In the meantime, we propose some recommenda-

tions (Table 2) to address the biases currently limiting fish passage.

We emphasize in the first instance the need to avoid creating new

barriers. New structures should be planned in a catchment or regional

context and, where deemed necessary from a socioeconomic perspec-

tive, be built in a manner that avoids or minimizes impacts on fish

movements. We recognize that remediation of existing structures

can be more challenging owing to existing site constraints and

legacies, but we highlight the need for removal to become the go‐to

option and for a more holistic approach to finding solutions where

removal is not practicable.

13

4 | CONCLUSION

In river ecosystems, fragmentation is a key driver of the Anthropocene

biodiversity crisis (Meybeck, 2003), raising alarm bells in the midst of a

global boom in dam building (Zarfl et al., 2015). Paradoxically, because

biodiversity and ecosystem functioning are inextricably linked, river

basin development aimed at supporting food, energy and water secu-

rity may actually be having the opposite effect. The uncritical applica-

tion of fishway technology has traditionally been the measure of

choice to mitigate connectivity losses, but it is increasingly seen as a

technology in decline. As is typically the case when a solution is not

working, the reasons why lie in its historical development. Early

fishways were conceived in response to the collapse of salmonid 1

stocks caused by a proliferation of migration barriers in northern

Europe. The migratory characteristics of salmonid species meant that

application of traditional, deterministic engineering approaches came

to dominate, specifically focusing on upstream migration. With the

realization that connectivity is important for taxa other than salmo-

nids, and the sharp increase in dam building outside of the temperate

Northern Hemisphere, came the erroneous assumption that salmonid‐

type fishways would work everywhere for all species. Evidence to the

contrary is now overwhelming but, as is usual with a paradigm shift,

the response lags behind. However, the debate is rapidly intensifying,

supported by the emergence of revised thinking, particularly from

outside the temperate Northern Hemisphere, and by increasingly

interdisciplinary training of practitioners. We have attempted to

contribute to this debate in the hope that continued discourse will

lead to better conservation of fish biodiversity in the near future.

We have highlighted examples that we believe represent progress

and proposed guiding principles for helping to advance the fish

passage discipline. However, if we fail to address these issues, we will

never reverse the loss.

ACKNOWLEDGEMENTS

This contribution was funded by the European Union AMBER project

(Adaptive Management of Barriers in European Rivers, #689682), the

Danish Fishing License Funds and the New Zealand Ministry for

Business, Innovation and Employment contract C01X1615. It was

further supported by the European Commission through the Marie

Sklodowska‐Curie action, ‘Knowledge Exchange for Efficient Passage

of Fish in the Southern Hemisphere’ (RISE‐2015‐690857‐KEEPFISH).

ORCID

Kim Birnie‐Gauvin http://orcid.org/0000-0001-9242-0560

Paul Franklin http://orcid.org/0000-0002-7800-7259

Martin Wilkes http://orcid.org/0000-0002-2377-3124

Kim Aarestrup http://orcid.org/0000-0001-8521-6270

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How to cite this article: Birnie‐Gauvin K, Franklin P, Wilkes

M, Aarestrup K. Moving beyond fitting fish into equations:

Progressing the fish passage debate in the Anthropocene.

Aquatic Conserv: Mar Freshw Ecosyst. 2018;1–11. https://doi.

org/10.1002/aqc.2946

It always seems impossible until it’s done - Nelson Mandela

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Technical University of Denmark

DTU AquaKemitorvetDK-2800 Kgs. Lyngby

www.aqua.dtu.dk

“The current scale of biodiversity loss in fresh waters is now so rapid that we consider it an invisible tragedy – hidden beneath the water surface – that attracts little public, political or scientific interest.” Reid et al. 2019