UNIVERSIDADE FEDERAL DO RIO GRANDE DO NORTE CENTRO DE BIOCIÊNCIAS
DEPARTAMENTO DE ECOLOGIA
PROGRAMA DE PÓS-GRADUAÇÃO EM ECOLOGIA
GRADIENTES EXPERIMENTAIS DE DIVERSIDADE INFLUENCIANDO HERBIVORIA E INTERAÇÕES INSETO-PLANTA EM COMUNIDADES VEGETAIS
RAFAEL DOMINGOS DE OLIVEIRA
NATAL 2019
2 RAFAEL DOMINGOS DE OLIVEIRA
GRADIENTES EXPERIMENTAIS DE DIVERSIDADE INFLUENCIANDO HERBIVORIA E INTERAÇÕES INSETO-PLANTA EM COMUNIDADES VEGETAIS
Tese de Doutorado apresentada ao Programa de Pós-Graduação em Ecologia, da Universidade Federal do Rio Grande do Norte - UFRN, como requisito para obtenção do título de Doutor em Ecologia.
Orientadora: Dra. Gislene Ganade (UFRN).
O presente trabalho foi realizado com apoio da Coordenação de Aperfeiçoamento de Pessoal de Nível Superior - Brasil (CAPES) - Código de Financiamento 001.
NATAL 2019
Universidade Federal do Rio Grande do Norte - UFRN Sistema de Bibliotecas - SISBI
Catalogação de Publicação na Fonte. UFRN - Biblioteca Setorial Prof. Leopoldo Nelson - Centro de Biociências - CB
Oliveira, Rafael Domingos de.
Gradientes experimentais de diversidade influenciando
herbivoria e interações inseto-planta em comunidades vegetais / Rafael Domingos de Oliveira. - Natal, 2019.
95 f.: il.
Tese (Doutorado) - Universidade Federal do Rio Grande do Norte. Centro de Biociências. Programa de Pós-graduação em Ecologia.
Orientadora: Profa. Dra. Gislene Maria da Silva Ganade.
1. Biodiversidade e funcionamento de ecossistemas - Tese. 2. Caatinga - Tese. 3. Restauração - Tese. 4. Herbivoria - Tese. 5. Redes de interação - Tese. 6. BrazilDry Experiment - Tese. I. Ganade, Gislene Maria da Silva. II. Universidade Federal do Rio Grande do Norte. III. Título.
RN/UF/BSE-CB CDU 574.1
3 RAFAEL DOMINGOS DE OLIVEIRA
GRADIENTES EXPERIMENTAIS DE DIVERSIDADE INFLUENCIANDO HERBIVORIA E INTERAÇÕES INSETO-PLANTA EM COMUNIDADES VEGETAIS
BANCA EXAMINADORA
______________________________________________ PROFA.DRA.GISLENE GANADE
Orientadora / Departamento de Ecologia
Universidade Federal do Rio Grande do Norte - UFRN
______________________________________________ PROF.DR.MÁRCIO ZIKÁN CARDOSO
Departamento de Ecologia
Universidade Federal do Rio Grande do Norte - UFRN
______________________________________________ DRA.MARÍLIA BRUZZI LION
Departamento de Ecologia
Universidade Federal do Rio Grande do Norte - UFRN
______________________________________________ DR.GUILHERME GERHARDT MAZZOCHINI
Instituto de Biologia / Departamento de Biologia Vegetal Universidade Estadual de Campinas - UNICAMP
______________________________________________ DR.LEONARDO HENRIQUE TEIXEIRA PINTO
Chair of Restoration Ecology
Department of Ecology and Ecosystem Management Technical University of Munich - TUM
4 “E esse jeito de deixar sempre de lado a certeza E arriscar tudo de novo com paixão Andar caminho errado pela simples alegria de ser.”
5 AGRADECIMENTOS
Todo o período do doutorado foi marcado por muito aprendizado e crescimento, tanto profissional, quanto pessoal. Muitas coisas aconteceram - várias boas, outras nem tanto, mas todas foram muito importante para minha formação. O saldo disso tudo é bastante positivo e, nesse momento, só tenho a agradecer a todos que contribuíram de alguma forma e fizeram parte dessa caminhada junto comigo.
Agradeço aos meus pais, Francisco e Liduina, por, mesmo sem saber exatamente a magnitude de toda essa jornada, me apoiaram, acolheram minhas decisões e, acima de tudo, sempre foram meus exemplos de caráter e integridade. Agradeço também ao meu irmão, Raniery, e ao meu sobrinho, Raul, por estarem ao meu lado sempre e me fazerem um bem danado!
À minha orientadora, Gislene Ganade. Muito obrigado pela confiança depositada, pelo total apoio, pelas grandes oportunidades e pelas valiosíssimas palavras e ensinamentos! Eu sempre falo de ti para todos com muito prazer e orgulho. Ter trabalhado no seu grupo e convivido contigo foi um dos maiores presentes desse doutorado. Gis, obrigado por ter remado por mim quando eu não consegui. Gratidão eterna!
Aos membros da banca, Dr. M árcio Zikán, pelo acompanhamento e contribuições desde o projeto inicial desta tese; ao Dr. Guilherme Mazzochini, pelas dicas em disciplinas, discussões e importantes contribuições em versões anteriores; ao Dr. Leonardo Teixeira e Dra. Marília Lion por terem gentilmente aceitado o convite para participar da banca e pelas pertinentes contribuições.
Ao professor Carlos Fonseca, por todas as contribuições durante o curso, na tese, nas disciplinas, nos momentos de descontração fora da Universidade e, principalmente, pelo imenso apoio na reta final deste trabalho!
À Dra. Carine Emer pela fundamental contribuição em um dos capítulos desse trabalho. Espero que façamos mais outras boas parcerias.
To Professor Wolfgang W. Weisser (TU Munich) for giving me the opportunity to work in one of the most important biodiversity experiment in the world, and for all the advices and contributions to the development of the study in Jena. I would like also to give special thanks to Dr. Sebastian Meyer (TU Munich) for the great support and supervision during my stay in Jena and Freising. Thank you for the time you patiently spent showing and teaching me data analyses and in R, and for the moments of informal chats on the “real life”. Thanks also to the other partners in Germany: Dr. Anne Ebeling (Uni-Jena), for making me feel as I was at “home”, providing me all the best office and field structure and for the best tips on the design and methods for insects’ samplings; Dr. Lionel Hertzog for the great help in the work and for the bike I used all the time in Jena! Thanks also to the technical
6 staff of the Institute of Ecology (Uni-Jena), specially to Maximilian Fraulob, und an alle Gärtner des Jena Experiment, vielen Dank!
To my flatmattes Lena Kath and Melissa Wich for the great time we had living together in Jena, sharing some bottles of wine and great typical german food they used to cook specially to make me know better their culture.
À Universidade Federal do Rio Grande do Norte e ao Programa de Pós-Graduação em Ecologia (PPGEco/UFRN), pela oportunidade da minha formação e auxílio financeiro para realização das pesquisas.
À Coordenação de Aperfeiçoamento de Pessoal de Nível Superior - Brasil (CAPES) – pelo apoio financeiro (Código de Financiamento 001), por meio de concessão da bolsa de doutorado; ao Conselho Nacional de Desenvolvimento Científico e Tecnológico (CNPq) pela concessão de bolsa de doutorado sanduíche no exterior.
This study was financed in part by the Coordenação de Aperfeiçoamento de Pessoal de Nível Superior - Brasil (CAPES) - Finance Code 001.
Ao Instituto Chico Mendes de Conservação da Biodiversidade (ICMBio), especialmente ao Chefe da Floresta Nacional de Açu (FLONA Açu), Mauro Guimarães, por concederem oportunidade de realização da pesquisa na unidade de conservação. Agradeço aos funcionários da FLONA, Irmão, Seu Chiquinho, Seu Luís e Moésio por todo o apoio.
Aos amigos do Laboratório de Ecologia da Restauração (LER) pela ajuda extrema nas coletas e realização dos experimentos e pelos vários momentos felizes que vivemos juntos! Muito obrigado mesmo! Agradeço à Marina Moura, Hudson, Luan, Aninha, Mayara, Ana Cleide, João, Gleide, Lara, Matthews e, especialmente, ao João Paulo pela grande ajuda em várias etapas do trabalho.
Aos meus brothers desse PPGEco Felipe Marinho, Marina Fagundes, Leonardo Teixeira e Paulo Marinho. Muita coisa boa a gente vivenciou... campos, viagens, cervejas, risadas, desabafos... Vocês são foda demais!
À minha querida amiga Nat por ser minha verdadeira companheira durante todo o tempo em Natal. Muitos cafés, muitas cervejas, muitas histórias. Obrigado pelas ajudas em campo, pelas discussões sobre nossos projetos, por todas as alegrias vividas na nossa casa e por ter aguentado e me segurado em todos os momentos de angústias. Realmente não sei o que teria sido de mim se não tivesse conseguido compartilhar com você tudo o que passou.
Ao Léo, que foi um grande companheiro, compartilhando experiências incríveis durante todo o tempo que estivemos juntos. Muito obrigado pela oportunidade de crescimento pessoal e profissional que tive ao seu lado e por todos os ensinamentos.
7 Ao Zoé e à Helô, amigos que me acolheram no RN, me levaram pra conhecer o Seridó, tomar caldo de cação em Caiçara do Norte e me deram o prazer de ser padrinho de casamento! Além dos baralhos, cachaças e comidinhas... sou fã de vocês!
A todos os meus amigos que torceram, se preocuparam em fases difíceis e que sempre estão ao meu lado: Bruno, Renata, Gaby, Ronaldo, Clarissa (bons momentos morando juntos!) e Maria Clara. Aos amigos Rafa, Lu, Lalá, Jamille e, especialmente, Cinthya por me abraçarem em Fortaleza e fazerem me sentir bem novamente no retorno às minhas raízes em um dos momentos mais delicados dessa empreitada toda.
8 SUMÁRIO RESUMO... 9 ABSTRACT………... 10 INTRODUÇÃO GERAL……….. 11 REFERÊNCIAS……… 14
Chapter I - Relationship between insect herbivory and plant diversity during dryland restoration... 16
Abstract…... 16
Introduction... 17
Material and methods... 19
Results... 25
Discussion... 30
References... 32
Supplementary material…... 37
Chapter II – Plant-herbivore network structure respond to plant diversity in restored tropical dryland ecosystem………..………. 40
Abstract…... 40
Introduction... 41
Material and methods... 42
Results... 46
Discussion... 55
References... 57
Supplementary material…... 59
Chapter III - Invertebrate herbivory along a plant diversity gradient in a trait-based experiment... 81
Abstract…... 81
Introduction... 82
Material and methods... 83
Results... 86
Discussion... 88
References... 89
9 RESUMO
Existem evidências de que a biodiversidade influencia diretamente ou está correlacionada com funções e serviços oferecidos e regulados pelos ecossistemas. Além disso, a diversidade de plantas pode afetar aspectos relacionados a níveis tróficos superiores, tais como abundância e diversidade de insetos herbívoros, os padrões de herbivoria e as interações multitróficas entre plantas, herbívoros e seus inimigos naturais. Desta forma, a presente tese aborda como a herbivoria por insetos pode ser afetada pela diversidade de espécies vegetais, pela diversidade de características funcionais das espécies vegetais, bem como investiga como a configuração da rede de interações entre insetos herbívoros e plantas pode mudar ao longo do tempo. O trabalho foi desenvolvido em experimentos de ampla escala na Caatinga Brasileira e em Campos de Gramíneas da Alemanha. Nesses experimentos comunidades vegetais foram construídas com diferentes níveis de diversidade de plantas. O capítulo 1 estuda como a herbivoria por insetos afeta a mortalidade de plantas e é afetada pela diversidade do plantio em um programa de restauração da Caatinga no semiárido brasileiro. Os resultados indicam que a diversidade de plantas nesse estágio inicial de restauração não afeta as taxas de herbivoria. Constatou-se também que a presença de insetos herbívoros não comprometeu o estabelecimento inicial de plantas durante a restauração de Caatinga, não sendo necessário o uso de qualquer inseticida para combater possíveis pragas que poderiam prejudicar a restauração. O capítulo 2 estuda as características da rede de interações entre insetos e plantas nos primeiros dois anos em experimento de restauração de Caatinga. O número de links foi maior em comunidades com desenvolvimento sucessional mais avançado, assim como elas apresentaram-se mais aninhadas. Existe a presença de espécies de insetos altamente generalistas nessas comunidades, assim como espécies de plantas que funcionam como hubs de interação na rede. O capítulo 3, realizado na Alemanha, aborda como a herbivoria por insetos pode ser influenciada tanto pela diversidade de plantas quanto pelos atributos morfo-funcionais dessas plantas. Foi constatado uma relação positiva entre a riqueza de plantas e as taxas de herbivoria por insetos. Além disso, as taxas de herbivoria são influenciadas pela sazonalidade. No entanto, a diversidade funcional e composicões de espécies formadas por diferentes traits funcionais (pools de espécies) não afetaram a herbivoria.
Palavras-chave: Biodiversidade e funcionamento de ecossistemas, Caatinga, restauração, herbivoria, redes de interação, Experimento BrazilDry, Experimento de Jena.
10 ABSTRACT
There is evidence that biodiversity directly influences or correlates with functions and services offered and regulated by ecosystems. In addition, plant diversity can affect aspects related to higher trophic levels, such as abundance and diversity of herbivore insects, herbivory patterns, and multitrophic interactions between plants, herbivores and their natural enemies. Therefore, this thesis examines how herbivory by insects can be affected by the diversity of plant species, by the diversity of functional characteristics of the plant species, as well as investigates how the configuration of the interaction networks between herbivore insects and plants may change over time. The study was performed in large-scale experiments in the Brazilian Caatinga and in Germany grasslands. In these experiments, plant communities were constructed with different levels of plant diversity. Chapter 1 studies how insect herbivory affects plant mortality and is affected by plant diversity in a Caatinga restoration program at the Brazilian semi-arid region. The results indicate that plant diversity at this early stage of restoration does not affect herbivory rates. It was also verified that the presence of herbivore insects did not jeopardize the initial establishment of plants during the Caatinga restoration, hence, it is not necessary to use any insecticide to combat possible pests that could harm the restoration. Chapter 2 studies the characteristics of the interaction network between insects and plants in the first two years of the Caatinga restoration experiment. The number of links was higher in older communities, as well as they seem more nested compared to young ones. The presence of highly generalist insect species are evident in these communities, and some plant species act as hubs of interaction in the networks. Chapter 3, conducted in Germany, discusses how insect herbivory can be influenced by both plant diversity and plant morpho-functional traits. A positive relation between plant richness and insect herbivory rates was found. Furthermore, herbivory rates are influenced by seasonality. However, functional diversity and species compositions grouped sorted by different functional traits (species pools) did not affect herbivory.
Keywords: Biodiversity and ecosystem functioning, Caatinga, restoration, herbivory, interaction networks, BrazilDry Experiment, Jena Experiment.
11 INTRODUÇÃO GERAL
A presente tese aborda como a herbivoria por insetos pode ser afetada pela diversidade de espécies vegetais, pela diversidade de características funcionais das espécies vegetais, bem como investiga a configuração da rede de interações entre insetos e plantas. O trabalho foi desenvolvido em experimentos de ampla escala na Caatinga Brasileira e em Campos de Gramíneas da Alemanha. Nesses experimentos comunidades vegetais foram construídas com diferentes níveis de diversidade de plantas. O capítulo 1 estuda como a herbivoria por insetos afeta a mortalidade de plantas e é afetada pela diversidade do plantio em um programa de restauração de Caatinga no semiárido brasileiro. O capítulo 2 estuda as características da rede de interações entre insetos e plantas transplantadas nos primeiros dois anos do experimento de restauração de Caatinga. O capítulo 3, realizado na Alemanha, aborda como a herbivoria por insetos pode ser influenciada tanto pela diversidade de plantas quanto pelos atributos morfo-funcionais dessas plantas.
Diversidade e estabilidade – história, mecanismos e hipóteses
Em termos históricos, a relação entre a diversidade de espécies e a estabilidade dos ecossistemas foi inicialmente verificada a partir de uma abordagem populacional (Odum 1953; Elton 1958) e de questões relacionadas à invasibilidade de espécies em comunidades naturais (McArthur 1955). Primeiramente, foi observada a dinâmica das populações de predadores quando esses possuem várias possíveis espécies de presas. Neste caso, o tamanho populacional desses predadores era pouco influenciado pelas flutuações causadas por variações ambientais da população de uma determinada espécie de presa, já que outras espécies de presas poderiam servir como alternativas alimentares. Além disso, foi observado que havia uma maior frequência e intensidade de ataques de pestes em sistemas agrícolas mais simples quando comparados aos sistemas naturais mais complexos. Esses mesmos padrões foram encontrados em comparações entre florestas boreais com menor diversidade de espécies e florestas tropicais em que a diversidade de espécies é maior (Odum 1953; Elton 1958). A relação entre diversidade e estabilidade dos sistemas foi colocada à prova na década de 1970. May (1973) propôs modelos matemáticos nos quais a existência de um número maior de espécies conferiu maior instabilidade do que os sistemas com menos espécies. Em outros casos, a relação entre a diversidade e a estabilidade variava, ora positivamente ora negativamente ou não apresentava efeito algum (Goodman 1975). McNaughton (1977) destacava a ausência de abordagens experimentais para constatação dessa relação. Portanto, a partir da década de 1990, David Tilman estabeleceu um experimento de longa duração para verificar a relação entre diversidade e estabilidade em comunidades vegetais (Cedar Creek Biodiversity Experiment). No mesmo período, Naeem et al.
12 (1994), realizaram experimentos em níveis multitróficos para entender como a redução da biodiversidade podem alterar a performance dos ecossistemas, causando uma reduçao na habilidade dos ecossistemas na absorção de CO2. Novos estudos experimentais começaram a surgir (BIODEPTH Experiment, BioCon Experiment, The Jena Experiment, TreeDivNet), com a finalidade de se entender, de uma forma geral, os efeitos da diversidade sobre a estabilidade de comunidades e ecossistemas (Tilman 2005). A partir desses experimentos foi constatado que a diversidade de espécies tende a apresentar uma correlação positiva com a estabilidade da comunidade vegetal (Tilman & Downing 1994; Tilman 1996; Tilman et al. 1996). Apesar da diversidade de espécies não exercer efeito ou mesmo diminuir a estabilidade de populações individuais, esta aumenta a estabilidade do ecossistema como um todo (Cottingham et al. 2001).
Estudos subsequentes mostraram que a estabilidade dos ecossistemas pode ser afetada por outros fatores que não a diversidade de espécies, como a topologia das redes tróficas (posicionamento das espécies em relação às trocas energéticas), a sensibilidade das espécies às flutuações ambientais, e as interações entre as espécies (Ives & Carpenter 2007). No entanto, a diversidade de espécies vem sendo bastante relacionada à estabilidade dos ecossistemas, em função de sua relativa facilidade de medição e manipulação, e por ser importante no debate sobre a perda mundial de biodiversidade (McCann 2000; Loreau et al. 2001). Apesar de a riqueza de espécies ser o aspecto de diversidade mais estudado e relacionado à estabilidade dos ecossistemas, as diversidades funcionais e filogenéticas também têm sido cada vez mais empregadas para detectar tais relações (Loreau & Mazancourt 2013; Venail et al. 2015). Cardinale et al. (2006) , observaram que a perda de espécies afeta o funcionamento de uma variedade de ecossistemas e, consequentemente, os processos ecológicos que controlam a abundância, biomassa e ditribuição de organismos. No entanto, a magnitude desses efeitos é determinada pela identidade das espécies que estão se extinguindo.
Existem várias evidências de que a biodiversidade influencia diretamente ou está fortemente relacionada com determinadas funções dos ecossistemas e, consequentemente, à estabilidade dos serviços oferecidos e regulados por eles (Cardinale et al. 2012). Mudanças na diversidade de organismos de um determinado ambiente podem provocar variações nos ciclos biogeoquímicos, na captura de CO2 e nitrogênio, (Chapin III et al. 2000), na produtividade dos ecossistemas (Tilman et
al. 1996), na resistência dos ecossistemas aos períodos de secas severas (Tilman & Downing 1994), e na diversidade, abundância e papéis desempenhados por organismos de níveis tróficos superiores (Duffy et al. 2007).
A diversidade influencia as funções ecossistêmicas por meio de mecanismos-chave, relacionados aos efeitos de complementaridade de espécies e aos efeitos de amostragem ou seleção (Loreau & Hector 2001). De acordo com o efeito de complementaridade, as espécies usam os recursos de diferentes maneiras, em diferentes situações (tempo, espaço, forma química), tornando a
13 exploração de determinado recurso mais eficiente, refletindo assim na melhoria das funções ecossistêmicas. Dessa forma, pode ser observado um efeito positivo da diversidade sobre a produtividade do ecossistema por meio de diferenciação de nicho ou da facilitação (Hector 1998; Cardinale et al. 2002). O efeito de seleção refere-se à presença de espécies funcionalmente eficientes dentro da comunidade, uma vez que comunidades mais diversas tem uma maior probabilidade de abrigar tais espécies (Loreau & Hector 2001). Assim, as funções ecossistêmicas são afetadas pela dominância de espécies com características particulares (maior eficiência na produção de biomassa, por exemplo). Portanto, algumas vezes o efeito da diversidade pode ser mascarado pelo efeito da espécie, ocorrendo o aumento do desempenho da comunidade além do esperado a partir do desempenho de espécies individuais (Huston 1997).
Existem algumas hipóteses formuladas para compreender a relação entre a diversidade de espécies e os processos ecossistêmicos (Naeem 1998). Segundo a hipótese da redundância (Walker 1992), as funções ecossistêmicas aumentam de acordo com o aumento do número de espécies, até certo ponto. Posteriormente, a adição de mais espécies é redundante e não exerce efeito adicional no funcionamento dos ecossistemas. A hipótese idiossincrática prevê a possibilidade de mudanças nas funções ecossistêmicas de acordo com o aumento ou diminuição do número de espécies, mas a direção das mudanças não é previsível (Naeem et al. 1995). De acordo com a hipótese do parafuso de Ehrlich e Ehrlich (1981), mesmo que de uma forma limitada, todas as espécies contribuem significativamente para a integridade do ecossistema como fazem os parafusos que sustentam a asa de uma aeronave. A redundância de espécies, desta forma, é importante porque a perda de uma única espécie não causaria o colapso de todo o sistema e somente a perda de um número grande de espécies levaria à perda de funções ecossistêmicas (Fonseca e Ganade 2001).
A presente tese investiga como a diversidade de espécies vegetais influencia processos de herbivoria e estruturação da comunidade de insetos que colonizam comunidades vegetais construídas experimentalmente em situação de campo. As várias hipóteses relacionadas a esse funcionamento específico são discutidas ao longo dos três capítulos apresentados. Os capítulos estão escritos em inglês para submissão em revistas de alcance internacional. O capítulo 1 será enviado para a revista Applied Vegetation Science, e investiga como a diversidade de árvores plantadas em um projeto de restauração da Caatinga influencia as taxas de herbivoria dos transplantes. O segundo capítulo será enviado à revista Oikos e investiga como as redes de interação entre espécies de insetos e plantas estão estruturadas nos diferentes níveis de diversidade de plantas na Caatinga. O terceiro capítulo foi realizado em campos de gramíneas da Alemanha e investiga como a diversidade funcional e a riqueza de espécies nas comunidades vegetais influencia o seu nível de herbivoria, esse capítulo será enviado à revista Perspectives in Ecology and Conservation.
14 REFERÊNCIAS
Cardinale, B. J., et al. 2012. Biodiversity loss and its impact on humanity. Nature 486:59-67. Cardinale, B. J., Palmer, M. A. & Collins, S. L. 2002. Species diversity enhances ecosystem functioning through interspecific facilitation. Nature 415:426-429.
Cardinale, B. J., Srivastava, D. S., Duffy, J. E., Wright, J. P., Downing, A., Sankaran, M. & Jouseau, C. 2006. Effects of biodiversity on the functioning of trophic groups and ecosystems. Nature 443:989-992.
Chapin III, S. F., Zavaleta, E. S., Eviner, V. T., Taylor, R. L., Vitousek, P. M., Reynolds, H. L., Hooper, D. U., Lavorel, S., Sala, O. E., Hobbie, S. E., Mack, M. C. & Diaz, S. 2000. Consequences of changing biodiversity. Nature 405:234-242.
Cottingham, K. L., Brown, B. L. & Lennon, J. T. 2001. Biodiversity may regulate the temporal variability of ecological systems. Ecology Letters 4:72-85.
Duffy, J. E.; Cardinale, B. J.; France, K. E.; McIntyre, P. B.; Thébault, E. & Loreau, M. 2007. The functional role of biodiversity in ecosystems: incorporating trophic complexity. Ecology Letters 10:522-538.
Ehrlich, P. & Ehrlich, A. 1981. Extinction, the Causes and Consequences of the Disappearance of Species.Victor Gollancz Ltd., London.
Elton, C. S. 1958. The Ecology of Invasions by Animals and Plants. Chapman & Hall, London. Fonseca, C. & Ganade, G. 2001. Species functional redundancy, random extinctions and the stability of ecosystems. Journal of Ecology 89: 118–125.
Goodman, D. The theory of diversity-stability relationships in ecology. The Quarterly Review of Biology 50:237-266.
Huston, M. A. 1997. Hidden treatments in ecological experiments: re-evaluating the ecosystem function of biodiversity. Oecologia 110:449-460.
Ives, A. R. & Carpenter, S. R. 2007. Stability and Diversity of Ecosystems. Science 317:58-62. Loreau, M. & Hector, A. 2001. Partitioning selection and complementarity in biodiversity experiments. Nature 412:72-76.
Loreau, M. & Mazancourt, C. 2013. Biodiversity and ecosystem stability: a synthesis of underlying mechanisms. Ecology Letters 16:106–115.
Loreau, M., Naeem, S., Inchausti P., Bengtsson, J., Grime J. P., Hector, A., Hooper, D. U., Huston, M. A., Raffaelli, D., Schmid, B., Tilman, D. & Wardle, D. A. 2001. Science 294:804-808.
MacArthur, R. H. 1955. Fluctuations of animal populations and a measure of community stability. Ecology 36:533–536.
15 May, R. M. 1973. Stability and complexity in model ecosystems. Princeton, NJ: Princenton
University Press.
McCann, K. S. The diversity–stability debate. Nature 405:228-233.
McNaughton, S. J. 1977. Diversity and stability of ecological communities: a comment on the role of empiricism in ecology. The American Naturalist 111:515-525.
Naeem, S. 1998. Species redundancy and ecosystem reliability. Conservation Biology 12:39-45. Naeem, S., Lindsey, J. T., Lawlers, S. P., Lawton, J. H. & Woodfin, R. M. 1995. Empirical
evidence that declining species diversity may alter the performance of terrestrial ecosystems. Phil. Trans. R. Soc. Lond. B 347:249-262.
Odum, E. P. Fundamentals of ecology (Saunders, Philadelphia, 1953).
Tilman, D. & Downing, J. A. 1994. Biodiversity and stability in grasslands. Nature 367:363–365. Tilman, D. 1996. Biodiversity: population versus ecosystem stability. Ecology 77:350-363. Tilman, D. 2005. Diversity and ecosystem services: does biodiversity loss matter? Proceedings of International Conference – Biodiversity Science and Governance. Paris. p. 95-102.
Tilman, D., Wedin, D. & Knops, J. 1996. Productivity and sustainability influenced by biodiversity in grassland ecosystems. Nature 379:718-720.
Venail, P., Gross, K., Oakley, T. H., Narwani, A., Allan, E., Flombaum, P. et al. 2015. Species richness, but not phylogenetic diversity, influences community biomass production and temporal stability in a re-examination of 16 grassland biodiversity studies. Functional Ecology 29:615-626. Walker, B.H. 1992. Biodiversity and ecological redundancy. Conservation Biology 6:18-23.
16 Chapter I
Relationship between insect herbivory and plant diversity during dryland restoration
Rafael Oliveira & Gislene Ganade
Programa de Pós-Graduação em Ecologia, Departamento de Ecologia, Centro de Biociências, Universidade Federal do Rio Grande do Norte, Av. Senador Salgado Filho, s/n, Lagoa Nova, 59072-970, Natal, Brazil.
Abstract
Drylands are one of the most susceptible areas to degradation due to climate factors and anthropogenic disturbances. Therefore, restoration practices in degraded tropical dry forests are urgent. Insect herbivory can be included among key factors affecting seedling and sapling survival in restoration programs. On the other hand, insect colonization is of extreme importance, since they are responsible for essential ecosystem functions. Plant diversity and complementarity may influence the abundance and diversity of herbivore insects and patterns of herbivory. This study aims to assess patterns of insect herbivory along a plant diversity gradient and the relation of herbivory damages to mortality in a restoration experiment of a degraded seasonally dry tropical forest from Northeastern Brazil. The results indicate that insect richness and abundance substantially increased from the first to the second year after restoration. Insect herbivory varied significantly between plant species, however plant diversity did not affect herbivory rates at this early stage of restoration. It was also verified that the presence of herbivore insects did not jeopardize the initial establishment of plants during Caatinga restoration, since there was no relationship between sapling mortality and insect herbivory. It indicates that the use of insecticide should not be recommended during early Caatinga restoration. Saplings in the experimental area played an important role for colonization of new higher trophic levels during Caatinga restoration. Colonization by these organisms will surely promote and enhance some functions and processes enabling natural ecosystem reestablishment over time.
Keywords: Caatinga restoration, sapling mortality, herbivory complementarity, herbivory damage, BrazilDry Experiment.
17 Introduction
Drylands are one of the most susceptible areas to degradation. Recent climate variations, such as increases in temperature, changes in rainfall dynamics, as well as environmental disturbances caused by human activities are important factors driving tropical dry forest degradation (Geist & Lambin 2004). The effects of these factors contribute to the high levels of degradation found in about 20% of drylands worldwide (MEA 2005a). This trend puts at risk essential ecosystem services for local human populations (MEA 2005b). Additionally, dryland degradation (desertification) is a highlighted topic in international organization programs, such as the United Nations Convention to Combat Desertification (UNCCD 2012). Given this scenario, restoration practices in drylands are of extreme urgency.
Restoration is a difficult task in arid and semi-arid lands (Bautista et al. 2009). In these environments, natural conditions have an extreme negative effect on restoration success of degraded areas (Paterno et al. 2016). Water stress is the major factor influencing these programs, since it is the main limiting resource for biological activities in drylands (Maestre et al. 2016). Scarce water availability in these areas does not allow constant irrigation to assist restoration processes. Moreover, long periods of drought jeopardize plant establishment and growth, hindering the development of the restored ecosystem. Common occurrences of fire and human activities such as logging and extensive grazing caused by domestic animals (mainly goat or cattle) are other factors influencing dryland degradation and restoration practices (Geist & Lambin 2004; Bautista et al. 2009; Marinho et al. 2016).
Insect herbivory can be included among key factors affecting seedling and sapling survival in restoration programs because great mortality rates are observed as a function of insect herbivory (Pritinnen et al. 2003; Sullivan 2003; Russel et al. 2010). In Steppe ecosystems, Cumberland et al. (2017) verified that the impact of grasshopper herbivory on native plant survival was prominent: they recovered 36% of plant mortality in the presence of these insects, compared to 2% of mortality in their absence. This pattern of native seedling mortality due to herbivory was also observed in a tropical dry secondary forest in Costa Rica (Gehardt 1993). Additionally, insect herbivory damage can strongly influence young seedling growth rates in restoration experiments from Atlantic forest (Massad et al. 2011) and semi-deciduous lowland forest (Plath et al. 2011). Furthermore, herbivory was associated with 50% decrease in native plant biomass in Steppe ecosystems (Cumberland et al. 2017). However, the use of pesticides in restoration programs is not advisable because many important ecosystem functions are played by insects, such as pollination, seed dispersion, predation, parasitism and detritivorism (Weisser & Siemann 2004).
18 Recent studies reveal that plant species diversity may influence the abundance and diversity of herbivore insects and patterns of herbivory (Ebeling et al. 2014; Hertzog et al. 2016; Hertzog 2017). This influence of plant diversity on higher trophic levels can be related to various factors (Duffy et al. 2007; Cardinale et al. 2011). For example, community plant biomass production, plant species richness, plant functional groups (Scherber et al. 2006; Loranger et al. 2014), habitat heterogeneity (Bommarco & Banks 2003; Abdala-Roberts et al. 2015) and degree of herbivore specialization are some factors influencing this relationship (Joshi et al. 2004). Furthermore, some studies have emphasized that herbivory response to plant diversity may depend on density (Kim & Underwood 2015; Kim 2017) and identity of plant species associated in mixtures (Orians & Björkman 2009; Castagneyrol et al. 2014; Underwood et al. 2014). Species identity may affect herbivore damage due to specific differences in plant quality and nutrient contents. Therefore, different plant species demonstrate positive or negative herbivory rates in response to plant diversity (Vehviläinen et al. 2007). Seasonality also drives these relationships through shifts in species-specific overall defensive chemistry, leaf contents (Riipi et al. 2002), and plant apparency in monocultures and mixtures (Moore & Francis 1991). Moreover, variation in insect herbivory can be due to physical characteristics of leaves, since low herbivory levels are observed in plant species with higher leaf hardness values (Dourado et al. 2016).
Effects of plant diversity on ecosystem functioning may also be due to complementarity effects. According to this approach, species use resources in different ways, therefore, in more diverse plant communities some ecosystem functions are improved by niche differentiation or facilitation (Hector 1998; Cardinale et al. 2002). Complementarity processes can also influence herbivory, since species associations may lead to increases or decreases of insect damage. Types of plant spatial associations may drive positive (associational susceptibility) or negative (associational resistance) relations (Barbosa et al., 2009; Underwood et al., 2014). Associational susceptibility consists in greater herbivory levels in plant species that are spatially associated with heterospecifics (Barbosa et al. 2009). The use of a wide range of food options by insect herbivores in high plant diversity allows them to enhance their performance and develop larger populations (Ebeling et al. 2018), increasing herbivory damage in these communities (Meyer et al. 2017). In associational resistance, insect herbivory on plant species decrease in taxonomically diverse plant communities (Tahvanainen & Root 1972; Barbosa et al. 2009). This may occur due to a dilution effect, because it is more difficult for insect herbivores, mainly specialists, to detect plumes of odor and find resources (Hambäck et al. 2014) or because specific resources for these consumers are less available (Castagneyrol et al. 2013). Moreover, higher richness and abundance of predators in more diverse plant communities lead to a suppression in herbivore abundance (Haddad et al. 2009; Letourneau et al. 2009) and, consequently, herbivory rates. Possible explanations for these associational effects on herbivory-diversity
19 relationship can be predicted by a set of mechanisms related to some hypotheses proposed in the literature (Table 1). Therefore, restoration experiments using replicated plant communities with different levels of diversity provide suitable scenarios to test these hypotheses (Blaisdell et al. 2015). This study aims to assess patterns of insect herbivory along a plant diversity gradient and the relation of herbivory damages to mortality in a restoration experiment of a degraded seasonally dry tropical forest from Northeastern Brazil. We addressed the following questions: 1) Do changes in plant diversity influence herbivory rates in restored communities? 2) Does plant diversity affect insect herbivory via complementarity? 3) Are there differences in herbivory rates among plant species identities? 4) Does herbivory affect plant survival in this semi-arid system?
Considering the degraded nature of the site, we predict that the insect herbivore community will consist mostly of generalist organisms. For that reason, our expectation is to detect higher herbivory levels in increasing plant species richness, since consumers would have a greater variety of resources availability to feed. Moreover, insect load and consequently herbivory damage are expected to be higher in more diverse communities, because consumers might be attracted by the greater biomass availability in these more productive systems. However, it is also possible that more diverse plant communities would present a higher variety of volatile secondary compounds that could inhibit overall herbivory (Scala et al. 2013).
Material and Methods
Experimental site
This study was carried out in the field site of the BrazilDry Experiment (5°32’17’’ S 36°57’49’’ W, 30 m a.s.l.), a restoration program established in July 2016 on 3.3 ha of a degraded area in a seasonally dry tropical forest (known as “Caatinga” vegetation), northeastern Brazil (Figure 1). The study area used to be an old grazing and agricultural field. However it is currently situated within a federal conservation area (Floresta Nacional de Açu, State of Rio Grande do Norte). The area is protected since 2001 but has no evidence of natural woody community regeneration. The conservation unit itself (432.52 ha in total) encompasses an important remnant of native Caatinga vegetation in a highly anthropized region. It consists predominantly of a preserved shrubby-arboreal community, although some degraded areas are present.
To assess the effects of plant diversity on insect colonization and herbivory, 135 13 x 8 m plots were established with five different levels of plant species richness: 1, 2, 4, 8, and 16. The field site was previously mowed and harrowed.
20 Table 1. Mechanisms related to hypotheses that might support the relationship between herbivory and plant diversity, according to the specialization level of herbivores. For each mechanism, some expectations are presented as possible effects of plant diversity on herbivory rates.
Hypothesis Specialization
level Mechanisms Expectations Bibliography
Resource-concentration
hypothesis
Specialist herbivores
Host finding by specialist herbivores can be hindered by the increasing number of non-host plants. The availability of the preferred plant species decreases due to reduced abundances of individual species in a
diverse plant community.
Decrease in specialist herbivores abundance and herbivory
with increasing plant species richness. Root (1973)
Dietary mixing hypothesis
Generalist herbivores
Generalist herbivores can directly profit from resource availability in higher plant diversity. They
have the capacity to deal with different toxins and can intake different nutrients, improving their reproductive success. They may also migrate from
more to less preferred plant species.
Increasing herbivore loads and damages with increasing plant diversity.
Bernays & Bright (1993); White & Whitham (2000); Unsicker et al. (2008) More individuals hypothesis Generalist herbivores
Resource availability to generalist herbivores increases with plant diversity due to higher productivity of high-diversity plant communities. It leads to increases in herbivore abundance and hence
in herbivore diversity.
Increases in abundance and diversity of consumers according to plant species richness promote increases in
herbivory rates. Srivastava & Lawton (1998) Productivity hypothesis Generalist herbivores
A higher overall resource level in diverse plant communities directly attracts more generalist
consumer species.
Positive relationship between herbivory and plant species
richness via increases in consumer diversity. Abrams (1995)
Enemies hypothesis
Generalist and specialist predators
High plant diversity limits overall herbivore abundance via shifts in predator communities. Predators and parasitoids of herbivores are more
abundant and diverse in species-rich plant communities.
Herbivory damages decrease with increasing diversity,
since preys (herbivores) are controlled by their predators. Root (1973)
Resource heterogeneity
hypothesis
Undetermined
A higher number of different resource in high-diversity communities increases the number of
consumer species.
If specialist herbivores dominate, increasing plant diversity increases the number of specialist consumer species and also community herbivory rates. If herbivory is mainly due
to generalists, then increasing herbivore abundance rather than diversity will increase herbivory.
Hutchinson (1959)
21 Figure 1. Experimental site location at the Açu National Forest, Rio Grande do Norte State, Brazil. Piató Lake consists in a fresh water natural reservoir that is strongly related to pluviometric dynamics of the region, therefore, subject to regular dry regims. Reservoir's soil and water are predominantly salty and it is not possible to use it for crop or planting irrigation.
In each plot, 32 saplings were planted 2 m away from each other and 0.5 m away from the plot edge (Figure 2). Plots were also 2 m apart. In the case of sapling mortality, new individuals were planted to maintain the same relative abundance of plant species in the experimental communities. Species distribution among plots was based on the Random Partition Design (Bell et al. 2009), where for each level of richness, species present in a plot are chosen from a pool without replacement, so that all plant species occur at the same densities in each richness level.
22 Figure 2. Schematic view of the full experimental design of the Caatinga restoration program with 155 13 m x 8 m plots with varying plant species richness (1, 2, 4, 8, 16 species and control plots). Plots are separated 2 m away from each other. The 32 individuals planted inside each plot were placed approximately 2 m apart. A total of 4704 individuals were planted. For herbivory assessment only 135 plots were sampled, all control plots and twelve of the 16 species plots were excluded. For insect assessment all plots apart from control plots were sampled.
Leaf sampling
This herbivory assessment was conducted during the two initial years of the restoration program. Leaf sampling was performed once in each plot at the end of the rainy season in June 2017 and June 2018, when individuals were ca. 1.6 and 2.6 years old, respectively. Sampling at the end of the rainy period represents the degree of damage accumulated during the growing season. This is so because during the rainy period water availability is ideal for plant growth, insect community is more abundant and active, therefore, herbivory might have its greatest impact. In Caatinga, most plant species shed their leaves during the dry season.
For herbivory assessment, leaves were sampled in plots with different levels of plant species richness (1, 2, 4, 8, and 16 species). To compare herbivory damages among all levels of plant species richness, two individuals of each species in each plot were selected. In order to reduce edge effects, individuals located in the central area of the plot were selected. When there were no leaves in plants located at this area or when individuals were recently replaced due to mortality, leaf sampling was
23 performed in the nearest co-specific plant. In the few cases where other suitable co-specific plant could not be found only one plant was sampled in the plot.
Two completely developed leaves were collected from each individual: the most basal leaf and the most apical leaf. Therefore, data on the accumulated damage made by insects from the beginning of the rainy season was expected to be gathered using older leaves. The younger leaves, which are generally preferred for consumption by herbivores, represented the damages made by a late season insect community. When there were no leaves at the upper branch, the most apical leaf at the branch immediately below was collected. When there were no leaves at the lower branch, the most basal leaf at the branch immediately above was collected. When there were no branches, leaves were collected from the main stem. Leaves folded by insects or spiders (as nesting or refuge sites) were not collected. We only collected two leaves per plant because some species had only few and large leaves (e.g., Pseudobombax marginatum and Handroanthus impetiginosus). By the end of the rainy season, some species were already shedding their leaves due to a decrease in the amount and occurrence of rainfalls. Sampled leaves were placed between wet sponges inside a cooler box with ice while in the field. This procedure was taken to avoid leaf shrinking due to drought and therefore provide perfect conditions for herbivory assessment in laboratory.
Herbivory measurements
All sampled leaves had their herbivory damage visually assessed and classified into four different types: chewing, rasping, mining, and sucking. In some cases, we used magnifying glasses to distinguish damages. Damaged leaf area was estimated separately for each damage type in eight percentage categories: 0, 1–5, 6–10, 11–25, 26–50, 51–75, 76–99, and 100% (Table 2). The range of each category was then transformed into a unique median value of insect herbivore damage, (i.e., 0, 3, 8, 18, 38, 63, 87.5, and 100%, respectively). The area damaged per leaf corresponds to the sum of the median values of herbivory of all damage types, since there was no overlap in damaged area between them. The herbivory damage per species represents the average damage value of all leaves for a given species (the same for the monospecific plots). For the species mixture plots, the total herbivory damage is the average damage value for all plant species present at the plot.
24 Table 2. Percentage categories for damaged leaf area and their respective median values used to calculate herbivory levels in plant species and plots of the Caatinga restoration program.
Categories of leaf area damaged Median of damage for category
0% 0% 1 – 5% 3% 6 – 10% 8% 11 – 25% 18% 26 – 50% 38% 51 – 75% 63% 76 – 99% 87.5% 100% 100% Insect sampling
Insect samplings were performed in May and June 2017 and May 2018 in all the 4704 woody individuals planted in the experimental restoration plots. In each plot every individual plant was inspected and insects spotted on the plant were manually collected. Insects were then placed in previously labeled vials for each different plant species inside a given plot. A piece of paper soaked in acetic acid ethyl ester (C4H8O2) was disposed inside each vial in order to quickly kill the collected
individuals. Insects were placed in a cooler box to appropriately conserve samples while they were taken to laboratory. They were then conserved in a 70% ethanol solution for posterior identification. Individuals were sorted to order, suborder or family when possible, and morphotyped using Operational Taxonomic Units (OTU). For future discussions, all individuals belonging to the order Orthoptera were considered as generalist herbivores, according to considerations on general patterns of feeding behaviour of insect herbivores in Bernays & Minkenberg (1997) and Bernays (1998).
Plant mortality assessment
In both years, 2017 and 2018, the restoration experiment was completely monitored for plant mortality. This assessment was performed for all the 4704 plants, i.e., for each individual of each plant species in each plot along all diversity gradient.
Data analysis
Herbivory complementarity was calculated according to a method adapted from Hector et al. (1998) for complementarity in plant biomass yielding. Firstly, one calculated the Relative Herbivory Complementarity of a plant species in a plot (RHC) from the following formula:
25 𝑅𝐻𝐶 = 𝐻𝑖𝑝
𝐻𝑖𝑚 𝑥 𝑁𝑖𝑝
where (Hip) represents the herbivory level of a plant species in a plot; (Him) represents the herbivory
level of the plant species while in monoculture; and (Nip) represents the number of individuals of the
plant species in the plot. The total value of Herbivory Complementarity for the plot (HC) was the sum of the RHC of all species in the plot:
𝐻𝐶 = ∑ 𝑅𝐻𝐶
When HC values are < 1, complementarity effects are smaller and plant species suffer less herbivory damage in species mixtures in relation to monocultures. When complementarity values are > 1, complementarity effects are larger and plant species suffer more herbivory damage in species mixtures than in monoculture.
To account for non-normality and heteroscedasticity of errors in the models, herbivory proportion data were arcsine square-root transformed, and herbivory complementarity values were log transformed.
To assess the effect of plant richness (log transformed) on the levels of herbivory and herbivory complementarity, we performed a Linear Mixed Effects analysis. We constructed a plant species matrix to ponder the possible relative contribution of each plant species present in the plots to the response of the ecosystem functions measured, i.e., herbivory and herbivory complementarity. Plant richness and plant species were stated as fixed effects, while species composition of the plots as a random effect. To assess if herbivory levels vary between plant species we also performed one-way ANOVA, followed by pairwise comparisons with a post-hoc Tukey test. To evaluate if plant mortality in the restored area was related to insect herbivory, we performed Linear Regressions where the number of dead plants in a plot was the response variable and insect herbivory was the explanatory variable. Analyses were performed in R 3.3.1, using the lmer function of the “lme4” package for the linear mixed effects analysis (Bates et al. 2015) and the functions aov, lm and TukeyHSD of the “stats” package for analysis of variance, linear regressions, and pair-wise comparisons (R Core Team 2016).
Results
Insect abundance was much higher in 2018 compared to the first year of the restoration program. In 2017, 769 insects belonging to nine orders were collected (Figure 3A). More than a half
26 of them (54.2%) were considered generalist insect herbivores. (Figure 3B). In 2018 this number substantially increased, reaching 3684 insects sampled, belonging to 11 orders (479.1% higher than the previous year; Figure 3A). From this total, 89.8% were considered generalist insect herbivores (Figure 3B).
Figure 3. Total abundance of insects (A), abundance of generalist herbivores (B) and average insect herbivory levels (C) sampled in 2017 and 2018 at the field site of the Caatinga restoration program (BrazilDry, Northeastern Brazil). Numbers over bars indicate the correspondent values for each response variable.
27 The overall levels of insect herbivory in the study area were 11.1% and 28.8% in 2017 and 2018, respectively (Figure 3C). These levels varied from 9.6% (in the 16 plant species mixtures) to 13.3% (in the 4 plant species mixtures) on average along the plant richness treatments in 2017 and 26.9% (in the 2 plant species mixtures) to 30.6% (in monocultures) in 2018. However, these variations in insect herbivory were not significantly different along the diversity gradient neither in 2017 (F(1,132)=2.73; p=0.11; Figure 4A) nor in 2018 (F(1,133)=0.0003; p=0.98; Figure 4B). Similar results
were found when the same analysis was performed for each plant species separately (Supplementary Material, Figure S1 and Figure S2, Table S1).
Figure 4. Levels of insect herbivory along a diversity gradient (1, 2, 4, 8 and 16 tree species) during the two first years after the experiment implementation (A=2017; B=2018) in a Caatinga dryland restoration program at northeast Brazil.
The overall values of herbivory complementarity were 1.19 and 1.02, in 2017 and 2018, respectively, varying on average from 1.06 (16 species mixtures) to 1.37 (4 species mixtures) in the first year (Figure 5A), and 0.92 (16 species mixtures) to 1.07 (8 species mixtures) in the second year (Figure 5B). The plant richness gradient did not affect herbivory complementarity in this early stage of restoration, neither in 2017 (F(1,85)=1.14; p=0.31) nor in 2018 (F(1,85)=0.56; p=0.47).
28 Figure 5. Average values of herbivory complementarity along a diversity gradient (1, 2, 4, 8 and 16 tree species) during the first two years (A=2017; B=2018) in a Caatinga dryland restoration program at northeast Brazil. Complementarity higher than one (dashed lines) represents higher herbivory in species mixtures than in monocultures. Error bars represent ± 1 standard error.
Insect herbivory levels varied significantly between plant species in both 2017 (F(15,341)=18.44;
p<0.001; Figure 6A) and 2018 (F(15,348)=11.53; p<0.001; Figure 6B). Combretum leprosum, Poincianella gardneriana, and Handroanthus impetiginosus were the species more damaged by insects (28.10%, 20.25% and 19.13%, respectively in 2017, and 38.94%, 37.20% and 42.47%, respectively in 2018). Piptadenia stipulacea also suffered high insect damage (37.90%) in 2018, although it presented the lowest proportion of herbivory (1.74%) in 2017. Compared to the other species at the same period, Commiphora leptophloeos presented one of the lowest levels of herbivory damage in both years (2.81% in 2017, and 11.58% in 2018). Sebastiania macrocarpa and Amburana cearensis were also species with low levels of insect damage (10.78% and 9.68%, respectively) in 2018, which were similar to what they have suffered in the previous year (5.96% and 11.73%, respectively).
29 Figure 6. Average herbivory levels observed per plant species in 2017 (A) and 2018 (B) for the whole plant species pool used in the Caatinga dryland restoration experiment in Northeastern Brazil. Means with different letters indicate statistically significant differences (Tukey’s HSD test, p<0.05). AmbCea = Amburana cearensis (Allemao) A.C.Sm. (Fabaceae); AnaCol = Anadenanthera colubrina (Vell.) Brenan (Fabaceae); AspPyr = Aspidosperma pyrifolium Mart. (Apocynaceae); BauChe = Bauhinia cheilantha (Bong.) Steud. (Fabaceae); CocVit = Cochlospermum vitifolium (Willd.) Spreng. (Bixaceae); ComLep = Commiphora leptophloeos (Mart.) J.B.Gillett (Burseraceae); CombLep = Combretum leprosum Mart. (Combretaceae); CynHas = Cynophalla hastata (Jacq.)
30 J.Presl (Capparaceae); HanImp = Handroanthus impetiginosus (Mart. ex DC.) Mattos (Bignoniaceae); LibFer = Libidibia ferrea (Mart. ex Tul.) L.P. Queiroz (Fabaceae); MimTen = Mimosa tenuiflora (Willd.) Poir. (Fabaceae); PipSti = Piptadenia stipulacea (Benth.) Ducke (Fabaceae); PoiGar = Poincianella gardneriana (Benth.) L.P. Queiroz (Fabaceae); PseMar = Pseudobombax marginatum (A.St.-Hil.) A.Robyns (Malvaceae); SebMac = Sebastiania macrocarpa Müll. Arg. (Euphorbiaceae); ZizJoa = Ziziphus joazeiro Mart. (Rhamnaceae).
The overall mortality rate in 2017 was 26.5%, varying from 3.7% in Commiphora leptophloeos to 58.2% in Amburana cearensis. In 2018, the overall mortality rate was 22.6%. Amburana cearensis was again the species with the highest number of dead individuals (61.2%). In turn, Piptadenia stipulacea presented the lowest mortality rate (3.4%) that year. There was no significant relationship between insect herbivory levels and plant mortality in 2017 (F(1,14)=0.67;
p=0.42; R2=0.04) and in 2018 (F(1,14)=4.19, p=0.06; R2=0.23).
Discussion
The levels of herbivory observed in the Caatinga restoration experiment are similar to those recorded in other dry tropical forests, especially for the herbivory assessment of 2017 (11.1% of insect herbivory), ranging from 6.7% to 17% (Bullock et al. 1995; Filip et al. 1995; Coley & Barone 1996). Bullock et al. (1995), however, found that many plant individuals from a dry tropical forest in Costa Rica showed moderate levels of damage (between 25% and 50%). Our data on herbivory in 2018 (28.8%) are similar to this research.
Plant diversity did not affect herbivory levels during the initial years of the restoration experiment. It was expected that insect herbivory would be higher in the more diverse plant mixture treatments, since generalist insects were abundant in the degraded area. When an insect community is dominated by generalist herbivores, these can benefit from the higher availability and variety of food resources characteristic of diverse plant communities (Bernays & Bright 1993; White & Whitham 2000; Unsicker et al. 2008). This pattern would lead to an increase in abundance and diversity of herbivores, as well as herbivory rates when plant diversity is higher (Srivastava & Lawton 1998). However, our results did not corroborate this expectation (related to the associational susceptibility hypotheses). We hypothesize that in the early years of community development plant biomass is still very similar between monocultures and species mixtures for an effect on insect dynamics to be detected.
Regarding complementarity effects, species mixtures showed slightly higher levels of complementarity (values >1) in 2017, but there were no differences in complementarity between the
31 distinct levels of diversity (from 2 to 16). This pattern indicates that in the first year after restoration, species mixtures in general tend to be more attractive to insect herbivores than monocultures. Therefore, complementarity could be an important mechanism by which diversity would increase levels of insect damage in more diverse plant communities. These positive complementarity effects were found in mixtures with more diverse genotype of oak species, where leaf consumption by generalist herbivores was increased due to the higher genetic diversity (Castagneyrol et al. 2012). In the second year, despite a non-significant effect, there was a tendency for not only a lower value of complementarity but also a general decrease in complementarity and lower herbivory levels in the 16 species mixtures in comparison to monocultures (complementarity < 1). Therefore, it is possible that this pattern becomes better established over the years while community matures, since relatively young communities (< 3 years) are still dominated by transient effects caused by an ephemeral assembly of invertebrate communities (Meyer et al. 2016).
It is important to point out that even though the insects used for assessment of insect density were found on the tree saplings they might not be using these plants as a food resource. It is likely that the herbivorous insects found at the experimental area could be using plants as shelter sites, because microclimatic conditions are quite harsh to the organisms inhabiting these regions (Joseph et al. 2016). Degraded areas in drylands have even worse conditions such as elevated temperatures, high radiation levels, intense wind blows, as well as low humidity. Even though insects have evolved to adapt to the dry environment, indigenous animals have to deal with these unusually harsh conditions of open disturbed areas (Thomas et al. 2014). Furthermore, Caatinga vegetation encompasses an extremely rich herbaceous flora that in disturbed lands occur as abundant pioneers, which is the case of our study site. Therefore, insects found at the study site could be feeding on the dense herbaceous layer established across the whole restored area and, therefore, the effect of plant diversity on herbivory levels may be masked until trees have acquired enough size to supress this thick herbaceous layer.
Plant species differed in the intensity they were eaten by herbivores which indicates that some plant species are more important as food or shelter for the insect herbivore community than others. In contrast to the McCreary and Tecklin (1994) observations in an oak restoration program, the great abundance of Orthoptera, mainly members of Acrididae (grasshoppers) and Proscopiidae (stick grasshoppers) families, was not a factor that have influenced plant mortality in Caatinga. Because higher herbivory levels did not result in higher plant mortality, we can conclude that the presence of herbivorous insects did not jeopardize early plant establishment during Caatinga restoration. This occurs because Caatinga plants have evolved to produce cheap leaf structures in this seasonal ecosystem, due to the fact that leaves will be surely lost when dry season arrives. Thus, small amounts of leaf loss would not influence plant survival. Indeed, small amounts of leaf loss could be even be
32 beneficial for plant survival because leaf loss by herbivory reduces leaf area for evapotranspiration preventing water loss in such a dry degraded area. These results indicate that the use of insecticide should not be recommended during early Caatinga restoration. Therefore, in an ecological restoration approach, this is an interesting result. In turn, we should highlight, the important role of the tree saplings planted in the experimental area for colonization of new higher trophic levels in the recent restored ecosystem of Caatinga vegetation. Even in the recent years of the restoration program establishment, several insects and even other animal taxa, as birds and amphibians, were observed using the experimental plants as hosts for nesting or refuge sites. Colonization by these organisms will surely promotes and enhances some functions and processes (e.g. nutrient cycling, pollination, seed dispersion, and decomposition), enabling reestablishment of ecosystem natural dynamics in time.
References
Abdala-Roberts, L., Mooney, K., Quijano-Medina, T., Campos-Navarrete, M., González-Moreno, A. & Parra-Tabla, V. 2015. Comparison of tree genotypic diversity and species diversity effects on different guilds of insect herbivores. Oikos 1: 1–9.
Abrams, P. 1995. Monotonic or unimodal diversity-productivity gradients: what does competition theory predict? Ecology 76(7): 2019–2027.
Barbosa, P., Hines, J., Kaplan, I., Martinson, H., Szczepaniec, A. & Szendrei, Z. 2009. Associational resistance and associational susceptibility: having right or wrong neighbors. The Annual Review of Ecology, Evolution, and Systematics 40: 1–20.
Bates, D., Maechler, M., Bolker, B. & Walker, S. 2015. Fitting linear mixed-effects models using lme4. Journal of Statistical Software 67(1): 1-48.
Bautista, S., Aronson, J. & Vallejo, V. 2009. Land restoration to combat desertification: innovative approaches, quality control and project evaluation. Fundación Centro de Estudios Ambientales del Mediterráneo – CEAM. Valencia, Spain.
Bell, T., Lilley, A., Hector, A., Schmid, B., King, L. & Newman, J. 2009. A linear model method for biodiversity–ecosystem functioning experiments. The American Naturalist 174(6): 836–849.
Bernays, E.A. 1998. Evolution of feeding behavior in insect herbivores. BioScience 48(1): 35-44. Bernays, E.A. & Bright, K. 1993. Mechanisms of dietary mixing in grasshoppers: a review. Comparative Biochemistry and Physiology 104(1) 125–131.
Bernays, E.A. & Minkenberg, O.P.J.M. 1997. Insect herbivores: different reasons for being a generalist. Ecology 78(4): 1157-1169.
Blaisdell, G., Roy, B., Pfeifer-Meister, L. & Bridgham, S. 2015. An exploration of hypotheses that explain herbivore and pathogen attack in restored plant communities. PLoS ONE 10(2): 1–16.
33 Bommarco, R. & Banks, J. 2003. Scale as modifier in vegetation diversity experiments: effects on herbivores and predators. Oikos 102(2): 440–448.
Bullock, S.H., Mooney, H.A. & Medina, E. Seasonally dry tropical forests. Cambridge University Press. 1995.
Cardinale, B., Matulich, K., Hooper, D., Byrnes, J., Duffy, E., Gamfeldt, L., Balvanera, P., O’Connor, M. & Gonzalez, A. 2011. The functional role of producer diversity in ecosystems. American Journal of Botany 98(3): 572–592.
Cardinale, B., Palmer, M. & Collins, S. 2002. Species diversity enhances ecosystem functioning through interspecific facilitation. Nature 415: 426–429.
Castagneyrol, B., Giffard, B., Péré, C. & Jactel, H. 2013. Plant apparency, an overlooked driver of associational resistance to insect herbivory. Journal of Ecology 101: 418–429.
Castagneyrol, B., Jactel, H., Vacher, C., Brockerhoff, E. & Koricheva, J. 2014. Effects of plant phylogenetic diversity on herbivory depend on herbivore specialization. Journal of Applied Ecology 51: 134–141.
Castagneyrol, B., Lagache, L., Giffard, B., Kremer, A. & Jactel, H. 2012. Genetic diversity increases insect herbivory on oak saplings. PLoS ONE 7(8): e44247.doi:10.1371/journal.pone.0044247. Coley, P.D. & Barone, J.A. 1996. Herbivory and plant defenses in tropical forests. Annual Review of Ecology and Systematics, 27: 305-335.
Cumberland, C., Jonas, J. & Paschke, M. 2017. Impact of grasshoppers and an invasive grass on establishment and initial growth of restoration plant species. Restoration Ecology 25(3): 385–395. Dourado, A., Sá-Neto, R., Gualberto, S. & Corrêa, M. 2016. Herbivoria e características foliares em seis espécies de plantas da Caatinga do nordeste brasileiro. Brazilian Journal of Biosciences 14(3): 145–151.
Duffy, J., Cardinale, B., France, K., McIntyre, P., Thébault, E. & Loreau, M. 2007. The functional role of biodiversity in ecosystems: incorporating trophic complexity. Ecology Letters 10: 522–538. Ebeling, A., Hines, J., Hertzog, L., Lange, M., Meyer, S., Simons, N. & Weisser, W. 2018. Plant diversity effects on arthropods and arthropod-dependent ecosystem functions in a biodiversity experiment. Basic and Applied Ecology 26: 50–63.
Ebeling, A., Meyer, S., Abbas, M., Eisenhauer, N., Hillebrand, H., Lange, M., Scherber, C., Vogel, A., Weigelt, A. & Weisser, W. 2014. Plant diversity impacts decomposition and herbivory via changes in aboveground arthropods. PLoS ONE 9(9): 1–8.
Filip, V., Dirzo, R., Maass, J.M. & Sarukhan, J. 1995. Within- and among-year variation in the levels of herbivory on the foliage of trees from a Mexican tropical deciduous forest. Biotropica, 27(1): 78-86.
Gerhardt, K. 1993. Tree seedling development in tropical dry abandoned pasture and secondary forest in Costa Rica. Journal of Vegetation Science 4: 95-102.
Geist, H. & Lambin, E. 2004. Dynamic causal patterns of desertification. BioScience 54(9): 817-829. Haddad, N., Crutsinger, G., Gross, K., Haarstad, J., Knops, J. & Tilman, D. 2009. Plant species loss decreases arthropod diversity and shifts trophic structure. Ecology Letters 12: 1029–1039.
