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UNIVERSIDADE FEDERAL DO RIO GRANDE DO SUL
INSTITUTO DE BIOCIÊNCIAS
PROGRAMA DE PÓS-GRADUAÇÃO EM ECOLOGIA
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André Frainer Barbosa
.
PORTO ALEGRE, ABRIL DE 2008
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UNIVERSIDADE FEDERAL DO RIO GRANDE DO SUL
INSTITUTO DE BIOCIÊNCIAS
PROGRAMA DE PÓS-GRADUAÇÃO EM ECOLOGIA
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André Frainer Barbosa
Orientador: Dr. Gilberto Gonçalves Rodrigues
PORTO ALEGRE, ABRIL DE 2008
Comissão
examinadora:
Prof. Dr. Albano Shwarzbold (UFRGS)
Prof. Dr. Adriano S. Melo (UFRGS)
Prof. Dr. Eduardo Mendes da Silva (UFBA)
Dissertação apresentada ao Programa de
Pós-Graduação em Ecologia, do Instituto
de Biocências da Universidade Federal do
Rio Grande do Sul, como parte dos
requisitos para obtenção do título de
Mestre em Ecologia
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À VÓ CARMEN LAIZOLA FRAINER
À VÓ ONELLA FANTINEL BARBOSA
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“A PRÁTICA É O CAMINHO DA VERDADE”
KARL MARX
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À Melina, agradeço pela paciência e companheirismo nesse tempo todo. Obrigado
pelo teu amor, minha namorada e esposa. À minha família: pai, mãe e meus irmãos Gus e
Duda. Com o apoio, o amor e a educação de vocês pude sonhar alto e pude, ao menos, tentar
alcançar grandes objetivos.
Aos companheiros de bancada, cuja amizade fez as longas horas de laboratório
passarem mais depressa: Verônica, Eduardo, Thiago, Cecília, Andréia e Luciana. Vocês
fazem parte dessa dissertação. Muito obrigado pela ajuda e apoio de vocês! Sou também
profundamente agradecido aos amigos que me ajudaram nas saídas-de-campo: Moisés, Lucas
e Andrés, além do Edu, Vero, Gil e Mê.
Aos meus grandes amigos que me acompanham desde a graduação (turma 2002/1 e
queridos(as) agregados(as)). Aos meus amigos que conheci na Pós ou cuja amizade se
aprofundou durante o Mestrado: Marcelo, Verinha, Fernando, Lê, Claudia, Haig They, Grazi,
Melina M. e Raul.
À todos os professores do Programa de Pós-graduação em Ecologia que me trouxeram
o conhecimento ecológico e deixaram grandes laços de amizade. À professora Inga L.
Veitenheimer Mendes, pela identificação dos exemplares de Mollusca citados nesse trabalho.
Aos funcionários do PPG-Ecologia (Silvana, em especial) e do Instituto de Biociências.
À FEPAM, em especial, à Ana Lúcia Rodrigues, Raquel Binotto e Maria Helena
Rodrigues pelo apoio logístico e pelos dados químicos e físicos utilizados nesse trabalho. Aos
motoristas Horácio Saraiva (in memoriam) e João Centena.
Aos moradores da Bacia Lajeado Grande, por transformarem as longas saídas-de-
campo em agradáveis visitas a amigos.
À CAPES pela bolsa de mestrado e ao povo brasileiro que mantém essa Universidade
pública, e cada vez mais popular, com qualidade.
Ao Gilberto Gonçalves Rodrigues que, primeiramente, me apresentou as maravilhas
do mundo aquático e me guiou nessa longa trajetória até os processos ecológicos. Sua
amizade e confiança me permitiram atingir os meus maiores objetivos científicos e
acadêmicos. Muito obrigado.
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RESUMO...................................................................................................................................... 1
ABSTRACT................................................................................................................................... 2
APRESENTAÇÃO.......................................................................................................................... 3
INTRODUÇÃO............................................................................................................................... 7
OBJETIVOS ................................................................................................................................ 11
CAP 1 – Sediment accumulation and not agricultural gradients influences leaf breakdown.... 12
CAP 2 – Influence of hydrological factors on leaf decomposition in a neotropical stream:
consequences for the shredders ....................................................................................... 30
CAP 3 – Use of BMWP and ASPT on southern Brazil through Cerrado and Argentinean values
.......................................................................................................................................... 46
CONSIDERAÇÕES FINAIS............................................................................................................ 61
REFERÊNCIAS BIBLIOGRÁFICAS................................................................................................. 66
ANEXOS .................................................................................................................................... 73
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TABELA 1. Environmental variables used to determine the land use types: percentage of
canopy cover, riparian vegetation width and agriculture intensity. ................................. 17
TABELA 2. Breakdown rates expressed in k value (± standard error) from light, moderate and
heavy agriculture areas..................................................................................................... 19
TABELA 3. Analysis of variance for the breakdown rate among heavy, moderate and light
agriculture areas. .............................................................................................................. 19
TABELA 4. Chemical and physical variables (mean values ± standard deviance) for light,
moderate and heavy agriculture areas. ............................................................................ 21
TABELA 5. Analysis of variance between sediment accumulated sites. ................................... 21
TABELA 6. Sediment backward stepwise multiple regression analysis. .................................. 21
TABELA 7. Correlation between sediment components. .......................................................... 22
TABELA 8. Water discharge (m
3
/s) measured monthly between October 2004 and January
2005. ................................................................................................................................ 33
TABELA 9. Mean chemical variables (± sd) taken monthly from October 2004 to January 2005
on each sampling site. ...................................................................................................... 34
TABELA 10. Breakdown rates obtained from Ocotea puberula leaves in slow and fast current
velocity sites. ................................................................................................................... 35
TABELA 11. Analyses of Co-Variance (ANCOVA) for genera richness, macroinvertebrates
abundance and density. ................................................................................................... 36
TABELA 12. Analysis of variance (ANOVA) of predator and prey ratio between treatments
and during exposure time. ................................................................................................ 36
TABELA 13. Analysis of covariance among Shredders mean number, treatments and time (as
covariate). ........................................................................................................................ 38
TABELA 14. Shredders richness ............................................................................................... 38
TABELA 15. Geographic coordinates of the sampling sites used for the biological assessment
.......................................................................................................................................... 50
vii
TABELA 16. BMWP values from three distinct regions: Brazilian Cerrado and Argentinean
Patagonia.. ........................................................................................................................ 52
TABELA 17. Analysis of Variance (ANOVA) for the BMWP values from Argentinean
Patagonia and Brazilian Cerrado...................................................................................... 52
TABELA 18. Analysis of Variance (ANOVA) for the ASPT values from Argentinean
Patagonia and Brazilian Cerrado...................................................................................... 53
TABELA 19. Permutation test for CCA under direct model...................................................... 53
TABELA 20. Tabela filogenética do total de amostras realizadas na Bacia do Lajeado Grande,
de novembro de 2004 a julho de 2005. ............................................................................73
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FIGURA 1. PCA for the environmental variables...................................................................... 20
FIGU RA 2. Sediment backward stepwise multiple regression significant term.. ......................22
FIGURA 3. Analysis of covariance (ANCOVA) for shredders density between treatments (slow
and fast water flow) and during time (as covariate). ....................................................... 37
FIGURA 4. Analysis of covariance (ANCOVA) for Shredders mean number between
treatments and during time (as covariate). ..................................................................... 37
FIGURA 5. Lajeado Grande catchment, southern Brazil. ......................................................... 51
FIGURA 6. ASPT values for sampling sites from Cerrado and Patagonia. .............................. 54
FIGURA 7. BMWP values for sampling sites from Cerrado and Patagonia. ............................ 55
FIGURA 8. CCA obtained for invertebrates` families and chemical and physical variables. ... 56
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O aporte de material foliar em sistemas lóticos de pequeno e médio porte é
responsável pela maior parte da energia e matéria que entra nesses ambientes. Esse material
passa por um processo de decomposição natural, o qual é afetado por diversos fatores, como
pela presença de microorganismos decompositores, macroinvertebrados fragmentadores ou
devido às condições físicas e químicas da água. Porém, enquanto que em regiões temperadas
os organismos retalhadores são reconhecidamente importantes para o processamento do
material foliar, em regões tropicais e sub-tropicais esses organismos parecem não participar
tão ativamente da decomposição foliar. Nesse trabalho, estudamos áreas influenciadas por
diferentes graus de antropização, objetivando verificar a influência de atividades agrícolas na
decomposição foliar, bem como a relação dos invertebrados retalhadores neste processo de
decomposição. Procuramos, também, avaliar a validade do uso de índices biológicos já
existentes para países da América Latina na detecção de impacto ambiental na bacia
hidrográfica em estudo. Os experimentos de decomposição foram realizados na Bacia
Hidrográfica Lajeado Grande, noroeste do Rio Grande do Sul. Nessa bacia hidrográfica, oito
sítios amostrais foram selecionados e tiveram suas características físicas e químicas
mensuradas. Em cada sítio, bolsas-de-folhiço contendo 4g de Ocotea puberula (Lauraceae)
foram expostas e retiradas mensalmente até o fim da decomposição foliar. Verificou-se que
constituintes físicos dos trechos amostrados, como frações da fácie areia, influenciam a
decomposição foliar negativamente (p = 0,0022), enquanto que o gradiente de uso da terra
analisado não demonstrou afetar a decomposição foliar (p = 0,3328). Por outro lado, a vazão,
característica desse sistema de corredeiras, mostrou-se responsável pela ação de disruptura do
material foliar com maior intensidade do que a ação de macroinvertebrados bentônicos. Os
valores de densidade, abundância e riqueza dos macroinvertebrados fragmentadores foram
afetados negativamente pela vazão da água. Esses organismos apresentaram, assim, maior
abundância e densidade nos tratamentos com menor vazão d’água e com a decomposição
foliar mais lenta (p = 0,0019). Estes resultados reforçam outros estudos realizados em
sistemas neotropicais que atestam que a decomposição foliar nessa região sofre maior
influência das variáveis físicas do que da atividade dos macroinvertebrados aquáticos, ao
menos no que se refere aos invertebrados retalhadores. O uso dos organismos relacionados ao
processo de decomposição foliar, na aplicabilidade de índices biológicos, também necessita
de estudos mais específicos, uma vez que os índices existentes para a América Latina
contemplam poucas regiões do continente.
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Litter fall in small and medium sized lotic systems is responsible for most of the
energy and matter input into these ecosystems. After the input of this material, it will be
decomposed by several factors, such as some physical water properties or the feeding
behavior of shredders and the presence of decomposer microorganisms. In temperate regions,
invertebrate shredders are well recognized to influence leaf decomposition. In tropical and
sub-tropical regions, however, these organisms seem not to present such important function.
In this work we studied some areas affected by human disturbance aiming to verify the
importance of land-use gradients to leaf decomposition as well as the role of shredders on this
ecological process. We also analyzed the applicability of rapid biological assessments on the
studied catchment using indexes already proposed for other Latin America regions. The study
was conducted on Lajeado Grande Basin, southern Brazil. There, eight sites were selected and
had some physical and chemical properties measured. In each site, four riffles were chosen
which were considered as replicates. In these riffles, litter-bags made of 4g Ocotea puberula
(Lauraceae) leaves were exposed and retrieved monthly until the end of the decomposition.
We verified that some physical constituents of the sites, as percentage of sand, influenced leaf
decomposition negatively (p = 0.0022), while the analyzed land-use gradient did not affect
leaf decomposition (p = 0.3328). On the other hand, water discharge - another physical
characteristic from these systems - was responsible for the leaf-litter breakdown with more
intensity than shredders were. Shredders’ density and abundance were affected negatively by
water discharge and presented greater numbers at the slow breakdown riffles (p = 0.0019).
These results are in agreement with studies from other neotropical systems, which suggest
that leaf decomposition in this region is more affected by physical variables than by biological
activity, at least for macroinvertebrate shredders. The use of the organisms that colonize leaf
material to the applicability of biological assessments also needs more studies, once the
existing indexes for their use in Latin America are appropriate for just a few regions.
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O estudo apresentado nesta dissertação teve início com a realização do projeto de
análise ambiental desenvolvido pela Fundação Estadual de Proteção ao Meio Ambiente do
Estado do Rio Grande do Sul (FEPAM-RS), juntamente com a Secretaria Estadual do Meio-
Ambiente (SEMA), intitulado Programa Nacional de Monitoramento Ambiental, fase 2
(PNMA-II) do Ministério do Meio Ambiente. Os resultados preliminares do projeto foram
compilados no Manual Técnico
1
da SEMA. Nesse estudo, aplicou-se o método de
decomposição de matéria orgânica, através do cálculo da sua taxa de decomposição, junto
com a análise da fauna associada ao material foliar exposto, como alternativa para a detecção
de fragilidades ambientais.
Ecossistemas aquáticos: o compartimento detrital
Estudos sobre a decomposição foliar em águas continentais remontam à década de 70,
quando diversos pesquisadores perceberam a importância da grande quantidade de material
orgânico proveniente da vegetação ripária em sistemas lóticos (e.g., Kaushik & Hynes, 1971,
Fisher & Likens, 1973; Petersen & Cummins, 1974), e também lênticos (e.g., Smock &
Stoneburner, 1980; Oerti, 1993), para a fauna aquática. Esse material orgânico formado em
sua maioria por folhas, mas também por outras partes vegetais, como galhos, troncos, e
diásporos (frutos e sementes), compõe grandemente a base energética dos trechos de rios de
pequena ordem em regiões temperadas (Webster et al., 1999).
O processo de decomposição foliar em sistemas lóticos inicia-se com o aporte de
folhas em ambientes aquáticos oriundos da vegetação ripária. Com a entrada do material
foliar na água, ocorre uma disruptura mecânica da folha juntamente com a dissolução de
diversos nutrientes presentes na mesma. Nessa primeira parte do processo pode ocorrer a
perda de cerca de 20% da biomassa original (Petersen & Cummins, 1974; Suberkropp et al.,
1976) em um período aproximado de 24 horas. Em seguida, há a colonização de fungos
hyphomicetes sobre esse material (Gessner et al., 2007), responsáveis pelo enriquecimento
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Rodrigues, G. G. ; Barbosa, A. F. 2006. Concepção Ecossistêmica para Avaliação da Qualidade da Água na
Bacia do Lajeado Grande. In: Niro Afonso Pieper. (Org.). Controle da Contaminação Ambiental Decorrente da
Suinocultura no Rio Grande do Sul: MANUAL TÉCNICO. Secretaria Estadual do Meio Ambiente. 1ed. Porto
Alegre, v. 2, p. 85-96.
4
das folhas com nutrientes retirados da água (Suberkrop & Klug, 1974; Bärlocher & Kendrick,
1974; Bärlocher, 1985; Suberkropp & Chauvet, 1995), ao mesmo tempo em que iniciam a
ação biológica de decomposição (Gessner & Chauvet, 1994; Gessner et al., 1999). Neste
mesmo período ocorre também a formação do biofilme bacteriano, o qual pode se constituir
em um importante componente da fase inicial de decomposição foliar (Hieber & Gessner,
2002; Buesing & Gessner, 2005). Após essa etapa de colonização por fungos e formação de
biofilme bacteriano, denominada também de fase de condicionamento, há a colonização por
macroinvertebrados bentônicos. Neste último grupo, a presença de organismos retalhadores
irá intensificar a decomposição foliar, sendo responsável pela degradação do restante da
biomassa foliar (Webster & Benfield, 1986).
A entrada de material foliar, geralmente, está associada a trechos superiores de rios
(até terceira ou quarta ordem), onde o sombreamento realizado pela vegetação ripária impede,
ou diminui consideravelmente, a produção autotrófica realizada pelo fitoplâncton e por algas
perifíticas. Por esse motivo, a produção heterotrófica torna-se fonte de matéria e energia
predominante para os sistemas lóticos de ordens inicias. Esse material alóctone se acumula,
principalmente, em zonas de remanso dos cursos d‘água de ordens iniciais ou nos trechos
intermediários, ou de transição, de rios. Em regiões temperadas, no entanto, a maior
velocidade de decomposição está relacionada aos trechos de corredeira, nos quais há maior
atividade microbiológica (Ferreira & Graça, 2006) e, dentre os macroinvertebrados, há maior
presença do tipo funcional denominado retalhador (Cummins et al., 1980; Kobayashi &
Kagaya, 2005).
Dessa forma, a disrupção do material foliar, ou material orgânico particulado grosso
(CPOM, do inglês coarse particulate organic matter), em partículas menores, ditas material
orgânico particulado fino (FPOM, do inglês fine particulate organic matter), une as
cabeceiras de rios com os trechos inferiores a partir da cadeia de detritos iniciada com a
deposição de material, oriundo, principalmente, da vegetação ripária (Vannote et al., 1980;
Webster et al., 1999). Vinculados à cadeia de detritos, a fauna de invertebrados aquáticos -
formada por organismos filtradores, coletores de partículas finas (denominados unicamente
coletores), raspadores de substrato (denominados unicamente raspadores) e predadores -
presentes em trechos intermediários e inferiores dos cursos d’água se beneficiarão e, em
última instância, dependerão do processo de decomposição foliar (Graça, 2001). Do mesmo
modo, a megafauna, ou macroconsumidores, - que compreende a íctiofauna, anurofauna, e
crustáceos - que se sucede nessa trama ecológica estará diretamente relacionada à cadeia de
5
detritos, iniciada nas cabeceiras dos cursos d’água (Vilella et al., 2005). Ressalta-se ainda que,
apesar da disruptura do material foliar ser melhor compreendida atualmente, a decomposição
de outras partes vegetais menos estudadas, como caules, ramos, e diásporos (frutos e
sementes), também apresenta importância para a cadeia de detritos em sistemas lóticos (Díez
et al., 2002; Harmon et al., 1986; Webster & Benfield, 1986).
Devido aos diversos compartimentos relacionados à decomposição foliar, essa
dissertação apresenta, primeiramente, uma introdução aos principais componentes atuantes no
processo de decomposição foliar. Após a introdução, seguem-se dois capítulos estruturados na
forma de artigos científicos, os quais discutem, cada um, parte dos compartimentos
apresentados na introdução e que, em linhas gerais, são os principais atuantes na
decomposição foliar.
O primeiro capítulo trata das características de mesoescala que afetam esse processo
ecológico. Inserem-se, nessa escala, os modos de uso da terra e as conseqüências desses usos
em escalas de menor abrangência (mesoescala), como a deposição de sedimentos devido à
supressão da mata ciliar. O segundo capítulo discute a influência dos macroinvertebrados do
tipo funcional retalhador no processo de decomposição foliar e os motivos pelo quais esses
organismos apresentam baixa importância em sistemas lóticos na região neotropical em
relação às regiões temperadas.
O terceiro capítulo, também estruturado como artigo científico, utiliza os valores de
abundância e riqueza de macroinvertebrados aquáticos presentes no material foliar para
determinar valores de qualidade da água, com base em índices biológicos pré-elaborados para
duas regiões distintas da América do Sul. Esse capítulo foi uma primeira tentativa de
estabelecer valores de qualidade de água para a região noroeste do Rio Grande do Sul e de
relacionar os macroinvertebrados presentes nessa região com valores químicos e físicos da
água, a partir de uma proposta dos órgãos ambientais do Estado do Rio Grande do Sul.
Os três artigos serão submetidos a diferentes revistas e, por esse motivo, estão aqui
formatados de maneira a manter a coerência estrutural da dissertação. Após o terceiro capítulo,
apresentam-se as considerações finais que pretendem vincular (i) os distúrbios físicos
oriundos do uso da terra para a decomposição foliar com as (ii) conseqüências dessas
alterações na estrutura e função da fauna aquática o que ocasiona, assim, uma (iii) diminuição
nos indicadores biológicos de qualidade da água. Estabelecer essa relação significa entender o
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quanto os processos de decomposição foliar são afetado por distúrbios antrópicos e, também,
compreender a validade do uso desse processo ecológico na detecção de alterações ambientais.
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Escalas de abrangência: variáveis abióticas da água (mesoescala) e uso da terra
(macroescala)
Diversos fatores atuam sobre o processo ecológicos de decomposição e influenciam na
velocidade com que o material foliar se degrada em sistemas lóticos (Leroy & Marks, 2006).
Entre os fatores físicos, inclui-se a velocidade da água como o mais importante, seguido de
efeitos de deposição de sedimentos (i.e., fácies de areia, silte e argila). Os picos de velocidade
da água podem acelerar o processo de decomposição consideravelmente, ao mesmo tempo em
que interferem na estrutura - composição, distribuição e abundância - da comunidade
bentônica (Arrington & Winemiller, 2006). A deposição de sedimentos finos, por sua vez,
atenua a intensidade desse processo ecológico em diferentes níveis de velocidade (Niyogi et
al. 2003; Sponseller & Benfield, 2001). Certos organismos com tipos funcionais específicos,
que atuam na decomposição (i.e., retalhadores), apresentam maior perda energética em
microhábitats onde haja alta velocidade d’água e formação predominante de areia, do que em
áreas com presença de cascalhos de maior diâmetro e de refúgios contra a correnteza (Franken
et al., 2006).
A maior incidência dos fatores físicos de correnteza e sedimentação pode estar
associada ao uso da terra no entorno dos cursos d’água, seja ele em função de atividades
agrícolas, seja devido à urbanização que intensifica consideravelmente essas variáveis.
Através das práticas agrícolas, o incremento de nutrientes como Nitrogênio e Fósforo pode
acelerar a decomposição foliar através do enriquecimento do material vegetal e, com isso,
proporcionar uma maior, ou mais rápida, colonização por fungos e/ou macroinvertebrados
aquáticos (Suberkropp & Chauvet, 1995; Robinson & Gessner, 2000; Gulis et al., 2006;
Bergfur et al., 2007). Por outro lado, o incremento de nutrientes pode, até mesmo, diminuir a
velocidade e/ou intensidade da decomposição por meio do impacto negativo sobre a fauna
bentônica (Baldy et al., 2007). Alguns estudos apontam, porém, para a inexistência de efeitos
da adição de nutrientes sobre o processo de decomposição. Raviraja et al. (1998) verificaram
que, apesar da alta poluição orgânica em um rio localizado na Índia setentrional, e do efeito
negativo dos poluentes sobre os fungos hyphomycetes, a taxa de decomposição não diferiu de
outro sistema aquático lótico sem a presença de poluentes, na mesma região.
8
Outro interferente que apresenta impacto negativo na fauna aquática e, por esse motivo,
no processo ecológico de decomposição subseqüente, é a aplicação de inseticidas em áreas
adjacentes aos cursos d’água (Cuffney et al., 1990; Wallace et al., 1995). Em áreas urbanas e
industriais, o acúmulo de metais pesados apresenta-se como fator de grande impacto no
retardamento do processo ecológico (Niyogi et al., 2001; Carlisle & Clements, 2005; Duarte
et al., 2008). A presença de focos pontuais de poluição (despejo de esgotos) tende a
enriquecer organicamente os sistemas aquáticos, de forma a ocasionar os mesmos efeitos que
o incremento de nutrientes em sistemas agrícolas exerce.
Ainda, o uso inadequado da terra, como o desmatamento da vegetação ciliar – entre os
mais comuns - irá diminuir o aporte de matéria orgânica para o sistema aquático (Niyogi et al.,
2004; Roberts et al., 2007), além de alterar a fonte de energia e matéria de um sistema
heterotrófico para autotrófico, devido à ausência de sombreamento exercido pela vegetação
marginal (Vannote et al., 1980). Áreas de desmatamento, mesmo quando realizadas dentro do
melhor critério possível, ou seja, sem corte de vegetação entre 30 e 100 metros de distância da
calha do rio, apresentam fortes impactos sobre a fauna aquática e, assim, sobre os processos
ecológicos de produção, consumo e decomposição foliar (Kreutzweiser et al., 2008). Ao
mesmo tempo em que a retirada da mata ciliar tende a modificar a fonte de energia e matéria
dentro do sistema aquático - de heterotrófico para autotrófico -, o aumento no aporte de
sedimentos devido à erosão e/ou lixiviação decorrentes do desmatamento também provocará
mudanças estruturais no curso d’água (Santmire & Leff, 2007). Essas mudanças tendem a
diminuir os valores obtidos em estudos de decomposição foliar através do soterramento do
material vegetal (Sponseler & Benfield, 2001). Devido à decomposição foliar ser afetada por
diversas alterações ambientais relacionadas ao uso da terra, esse processo foi proposto para
ser utilizado na indicação ambiental em cursos d’água (Gessner & Chauvet, 2002; Royer &
Minshall, 2003).
Macroinvertebrados aquáticos
Estudos em regiões temperadas demonstram que a presença de macroinvertebrados
alteram a velocidade de decomposição do material foliar (e.g., Webster et al., 1999) e, com
isso, agem na ciclagem de nutrientes dos ecossistemas aquáticos. Na região tropical e sub-
tropical os retalhadores strictu sensu (e.g., Plecoptera:Gripopterygidae) e também
macroconsumidores (e.g., Crustacea:Trichodactilydae), se destacam (Rosemond et al., 1998;
9
Vilella et al., 2005) em oposição a outros taxa descritos para as regiões temperadas (e.g.,
Crustacea:Gammaridae). Tem-se, contudo, pouca evidência do papel desempenhado por esses
organismos na decomposição foliar (Mathuriau & Chauvet, 2002; Moretti et al., 2007), em
especial na região neotropical.
Apesar do aporte de material foliar ser em grande quantidade nos sistemas aquáticos
tropicais e sub-tropicais, durante o ano todo (Nin et al., no prelo), dentre o grupo de
macroinvertebrados aquáticos, encontram-se poucos insetos verdadeiramente retalhadores
nessas regiões (Wantzen & Wagner, 2006). Um dos motivos dessa diferença em relação às
regiões temperadas aparece devido à grande quantidade de predadores presentes em sistemas
netropicais, os quais influenciam na dieta dos insetos, e fazem, assim, com que os organismos
detritívoros apresentem comportamento mais generalista para obtenção de alimentos (Covich,
1988; Wantzen & Wagner, 2006). Além disso, insetos aquáticos tropicais e sub-tropicais
apresentam ciclos de vida pronunciadamente mais curtos do que insetos de regiões
temperadas. Essa característica, juntamente com a maior predação, tende a levar os insetos a
possuirem uma dieta onívora, a fim de acelerarem seu desenvolvimento em insetos adultos.
Alguns autores (e.g., Graça, 2001) atestam que a natureza morfológica e química do
material foliar de espécies de regiões neotropicais, por apresentarem maior diversidade devido
às pressões ecológicas, possui, também, maior quantidade de compostos secundários
utilizados na defesa contra a herbivoria (Feeny, 1970; Grime et al., 1996; Poorte et al., 2004;
Wantzen et al., 2002; Santiago, 2007). Esses compostos, junto com características
morfológicas (e.g., tricomas) poderiam continuar atuando depois da entrada do material nos
sistemas aquáticos, de forma a impedir ou, ao menos, prejudicar a colonização das folhas por
macroinvertebrados aquáticos.
Outro ponto ainda, seriam as fortes chuvas tropicais que podem carrear o substrato dos
sistemas lóticos e suprimir a fonte de recursos alóctone que serviria para a alimentação dos
retalhadores (Winterbourn et al., 1981). Por este aspecto, organismos detritívoros não podem
depender de um único recurso natural, que, apesar de ter aporte constante em regiões tropicais
e sub-tropicais, é instável dentro do sistema aquático. Assim, acredita-se que a menor
abundância e riqueza de retalhadores verificados em regiões neotropicais (Wantzen & Wagner,
2001) façam com que a decomposição foliar seja influenciada mais diretamente pela ação de
microorganismos como fungos hyphomicetes, ou mesmo pela ação física da velocidade da
água.
10
Uso de macroinvertebrados no biomonitoramento
Os macroinvertebrados bentônicos são uma categoria ecológica amplamente utilizada
para a detecção de alterações ambientais (Rosenberg & Resh, 1993) devido ao fato de terem
suas características estruturais (e.g., ocorrência e abundância) e funcionais (e.g., presença de
grupos alimentares específicos) influenciadas pelas condições físicas, químicas e geoquímicas
do ambiente (Robinson & Minshall, 1986; Lake, 2000; Gafner & Robinson, 2007). Além
disso, esse grupo apresenta larga distribuição espacial e são encontrados em alta abundância
em ambientes lóticos e lênticos, o que faz deles organismos propícios para a coleta e estudo
em praticamente qualquer sistema aquático continental (Rosenberg & Resh, 1993).
A ampla distribuição e alta abundância de macroinvertebrados aquáticos, aliado ao
fato de eles serem identificados com razoável facilidade e de serem coletados e armazenados
com equipamentos simples, fez com que diversos índices e métricas para a detecção de
alterações ambientais pudessem ser elaborados a partir da ocorrência desse grupo de
organismos (Resh & Jackson, 1993). Em países como Inglaterra, Austrália e Estados Unidos,
os macroinvetebrados bentônicos são utilizados juntamente com análises químicas e físicas da
água para o monitoramento, análise e indicação da qualidade da água em sistemas naturais ou
antropicamente alterados (e.g., Armitage et al., 1983). Na América Latina e no Brasil, em
especial, esses índices estão sendo recém estudados e aplicados (Barbosa et al., 2001; Cota et
al., 2001). Por esse motivo, estudos sobre macroinvertebrados aquáticos que foquem o
biomonitoramento ambiental devem ser realizados no Brasil para, com isso, aumentar tanto o
conhecimento taxonômico pertinente, quanto o entendimento sobre a distribuição espacial dos
organismos e suas relações com os fatores ambientais.
11
O
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Objetivo geral
- Compreender a importância dos macroinvertebrados aquáticos na decomposição
foliar em um sistema lótico neotropical e compreender se esse processo ecológico pode ser
utilizado para verificar efeitos antropogênicos em um gradiente de alterações ambientais
Objetivos específicos
- Verificar se diferentes tipos de usos da terra afetam a decomposição foliar.
- Determinar a importância de variáveis bióticas e abióticas para o processo de
decomposição foliar em um sistema neotropical.
- Estabelecer índices biológicos de qualidade da água para os trechos estudados a
partir da fauna de macroinvertebrados associada ao folhiço.
12
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(Com Gilberto G. Rodrigues)
(Artigo submetido à Freshwater Biology, e em preparação para ser re-enviado.)
“
If I had a world of my own, everything would be
nonsense. Nothing would be what it is, because
everything would be what it isn't.”
Alice in wonderland,
Lewis Carroll
13
Abstract
Leaf-litter is a major resource that enters streams and is responsible for most of the
headwaters energy budget. Its decomposition is influenced by macroinvertebrate shredders
and fungi hyphomicetes and by chemical and physical water properties. Several studies have
tried to link leaf processing to anthropogenic disturbs. We aimed to verify whether land use
types would influence leaf decomposition. Eight stream reaches were selected in southern
Brazil and classified into three groups: heavy agriculture, moderate agriculture and light
agriculture. A few chemical parameters differed from heavy agriculture to moderate and light
agriculture. Our results showed that leaf breakdown rates were lower both in the heavy and in
the moderate agriculture sites in comparison to the light agriculture sites (p = 0.03). The main
cause for slowing down leaf breakdown rate was sediment accumulation in the heavy
agriculture sites. Leaf decomposition was positively related to percentage of sand texture (p =
0.002) and this texture was negatively related to silt and clay. The sites with lower
decomposition rates had more clay and silt on its sediment composition. These findings
suggest that leaf decomposition is influenced by agriculture status due to land use alterations,
as riparian vegetation removal, which can increase local sedimentation and imply negative
consequences to litter breakdown.
Key-words: ecological assessments, leaf breakdown, land use, sediments.
14
Introduction
The ecological importance of allocthonous material input and its decomposition in
canopy covered stream systems has been widely reported (Kaushik & Hynes, 1971; Vannote
et al., 1980; Gessner et al., 1999; Webster et al., 1999). According to Gessner and Chauvet
(2002), ecosystem processes are very useful when one wants to understand the health (sensu
Karr, 1999) of an ecosystem and determining litter breakdown is one potentially useful way to
achieve this goal (Bunn & Davies, 2000; Gessner & Chauvet, 2002, Royer & Minshal, 2003).
However, recent field tests of the idea have led to some apparently conflicting results. While
some authors have found a close relation between litter breakdown and measures of human
disturbance (Pascoal et al., 2003; Niyogi et al., 2003, Leroy & Marks, 2006), other authors
argued that it is difficult to relate decomposition rates to types of land use (Hagen et al., 2006,
Sponseller & Benfield, 2001).
Leaf breakdown involves several organisms, primarly aquatic hyphomycete fungi
(Suberkropp et al., 1983; Gessner & Chauvet, 1994; Baldy et al., 1995; Pascoal et al., 2005)
and benthic macroinvertebrates (Bird & Kaushik, 1992; Linklater, 1995; Rodrigues, 2001;
Gonçalves et al., 2006). Higher nutrient (N or P) input from agriculture may result in higher
microbial activity and higher breakdown rate (Suberkropp & Chauvet, 1995; Robinson &
Gessner, 2000; Ferreira et al., 2006). Elevated phosphorous concentrations have been related
to increased macroinvertebrate density, especially of shredders (Rosemond et al., 2001, Paul
et al., 2006), and may therefore positively affect decomposition rates. Agricultural activities
are also usually accompanied by removel of natural riparian vegetation and substitution by
crop plantations. This changing on riparian structure results in increased sedimentation on
streams (Townsend & Riley, 1999) which, in turn can decrease both microorganisms
(Santmire & Leff, 2007) and invertebrates density and richness (Zweig & Rabeni, 2001;
Chaves et al., 2005) and thus may slow down litter decomposition (Sponseler & Benfield,
2001, Niyogi et al., 2003).
In this study we used exponential leaf breakdown rates obtained as k values (Petersen
& Cummins, 1974) from eight stream sites along a small catchment (˜ 500 ha) in southern
Brazil to assess leaf breakdown response to environmental disturbances. The hypothesis was
that leaf decomposition rate will be negatively affected by the intensification of agriculture in
a region affected by livestock grazing and other intensive farming activities.
15
Material and Methods
Study site
Lajeado Grande Basin is located in southern Brazil, Rio Grande do Sul State and is in
the Broadleaf Subtropical Forest area. The area presented human activities linked specially to
pig farming, a well known non-point source water contaminant. Several small catchments
(usually around 90 km length each) present this economic activity in Rio Grande do Sul State,
and the Lajeado Grande Basin is among those with the highest density of pig growing farms.
The cacthment area of Lajeado Grande Basin is 525.38 km
2
. Stream depth is around
0.5 m in the headwaters and 4 m near the river mouth. Together with Rio Grande do Sul State
Environmental Agency (FEPAM) eight reaches were chosen to determine leaf breakdown
rates. Four of these reaches were located on the main stream channel, the Lajeado Grande,
and the other four reaches were located along its main tributary stream, the Lajeado Erval
Novo. Stream substrate was composed predominantly of gravel and sand while catchment’s
soil is predominantly made of fine particles (FEPAM, unpublished data).
Stream reaches located in the mean river are called LG after Lajeado Grande and
stream reaches located in the mean tributary are called LEN, after Lajeado Erval Novo. All
sites are numbered from one to four due to its location in the system: smaller numbers refer to
sites close to river mouth and higher numbers refer to sites located close to the headwaters
(Figure 1).
Leaf-bags
Ocotea puberula (Rich.) Nees (Lauraceae family) is a very common tree in northern
Rio Grande do Sul and occurs all over Brazil and mostly South America region, from the
French Guyana to northern Argentina. Leaves were collected from one single tree in order to
avoid possible leaves components differences between individuals and between eco-regions.
Leaves were then air-dried until necessary and weighed to the nearest 0.01g.
Leaves were weighed and packed into 10 mm mesh size ordinary nylon bags. Each
pack contained 4.0 g of O. puberula leaves (exact weight of each bag was recorded) and
labelled one-by-one with PVC marked tags. A total of 320 leaf-bags was made and divided
16
into eight groups corresponding to the eight study stream sites. For each reach 40 leaf-bags
were placed among four sample points fixed on the river banks (10 leaf-bags for each
sampling point). To maximize representativeness of within-site heterogeneity, main
environmental conditions in each reach (pools and riffles) were sampled.
Four replicates were collected at each site after the following periods of exposure: 15,
30, 60, 90, 120, 150 and 180 days. Litter-bags were gently removed from water using a
200µm mesh handy-net and immediately stored on previously labelled plastic bags. For
transport to the laboratory litter-bags were kept in a cooler. Leaf-bags were opened in
laboratory, washed over another 200µm mesh size screen with tap water for cleaning leaves
from associated material and for retaining the macroinvertebrate associated fauna for further
analyses. Leaves were oven-dried at 60 °C for 48h and weighed to the nearest 0.01g.
Four reaches were chosen to have the one-day exposure period. The mean weight
obtained between these reaches was used as initial leaf-bags weight. Breakdown rate was
measured by the k coefficient formula M
t
= M
o
e
-k.t
(Petersen & Cummins, 1974), where: M
t
=
mass at time t; M
o
= leaf initial mass; t = time in days; and k = exponential decay coefficient.
Figure 1. Lajeado Grande catchment in southern Brazil. Sampling sites are shown.
17
Chemical and physical variables
Total nitrogen, nitrate, phosphorous, dissolved oxygen, specific conductance, pH-
value, and water discharge were taken monthly from October 2004 to January 2005.
Environmental variables
During field procedures, surrounding environment features (agricultural intensity,
particle composition, riparian vegetation width and riparian canopy cover on the stream) and
characteristics of the channel (predominant substrate, maximum depth and channel width)
were taken. Agricultural intensity was measured as a modification of Hagen et al. (2006) site
description: 0 = no agricultural activity next to the stream; 1 = active agriculture next to the
stream but no livestock grazing; 2 = agriculture and livestock grazing reaching stream banks.
Surface refers to the mean ground composition. 1 = mainly gravel; 2 = gravel and silt; 3 =
mainly silt. Riparian vegetation width was measured as: 0 = zero to one meter; 1 = one to five
meters; and, 2 = five to ten meters vegetation width. Riparian canopy cover was measured as
percentage of stream shelter: 0 = 0%; 1 = 25%; and 2 = 50% (Table 1).
Table 1. Environmental variables used to determine the land use types: percentage of canopy cover,
riparian vegetation width and agriculture intensity.
Sediment process
During the experiment we observed that one of the light agriculture areas, namely LG2,
lacked on its riparian vegetation in one of the stream banks. This fact increased local
sedimentation over the leaf packs. Also, the presence of a small dam (one meter high)
downstream could help sediment accumulation backwards the dam localization. We thought
that sedimentation could be another factor influencing breakdown rate rather than only land-
Low agriculture Moderate agriculture Heavy agriculture
Discharge [m
3
/s] 0.53 – 10.32 0.32 – 7.34 0.0009 – 0.004
Channel width [meters] 6 – 10 6 – 10 1
Channel max. depth [meters] 0.4 – 2 0.35 – 3 0.2 – 0.5
Surface 1 – 3 2 3
Canopy cover 1 – 2 0 – 2 0 – 2
Riparian vegetation width 2 1 0
Intensive agricultural area 0 – 1 1 2
18
use type. We used sediment analysis from the river banks measured by FEPAM (data
unpublished) from October 2004.
Statistical analyses
Each water chemistry variable was tested by one-way ANOVA. The environmental
variables of the land-use classification were analyzed after vector transformation within
variables by standardizing by the range. A matrix of Bray-Curtis dissimilarity resemblance
measure between the sampling units was obtained. A Principal Component Analysis (PCA)
was then executed. These procedures were executed on the MULTIV software (version 2.3.20.
Copyright © Pillar, V. D. 2006).
Leaf breakdown was compared among land-use categories after ln(x+1)
transformation of remaining mass in an Analysis of Co-Variance test (ANCOVA), followed
by a posteriori Tukey’s honest significant difference (HSD). A linear regression model
between k values as response variables and the first PCA axis as independent variable was
executed. Also an one-way ANOVA between sites with high sedimentation levels and sites
with low or no sedimentation was done. We, then performed a backward stepwise multiple
regression analysis, in order to find which of the sediment components (sand, gravel, clay and
silt) would be influencing the breakdown rates. Sediment levels original data was as
percentual substrate cover. We transformed this data into arcsine square root (Zar, 1999).
Correlation between sediment components was verified. These tests were performed in the R-
Program (version 2.3.0. R Development Core Team, 2006).
Results
Leaf breakdown
Three levels of breakdown rate could be classified among the stream reaches fitting
Petersen & Cummins (1974) breakdown velocity classification: fast breakdown rates (LG1,
LG3 and LG4 sites, k ˜ 0.010 day
-1
), slow-moderate breakdown rates (LG 2, LEN 1 and LEN
2 sites, k ˜ 0.004 day
-1
) and slow breakdown rates (LEN 3 and LEN 4, k ˜ 0.003 day
-1
). Leaf
breakdown among agriculture land uses types (Table 2), however, did not differ statistically
19
(p = 0.33, Table 3). At LG1, leaves were totally decomposed after 90 days. At LG3 and LG4
there were no more leaves in litter-bags in 120 days, while LG2, LEN1, LEN2, LEN3 and
LEN4 decomposed only after 180 days.
Table 2. Breakdown rates expressed in k value (± standard error) from light,
moderate and heavy agriculture areas.
Land use gradient
After conducting the PCA, the first stream classification made had to be changed, as
LG3 moved to the moderate agriculture group (Figure 2). The first axis of the PCA
corresponds to 74.16% of total variation and the second axis corresponds to 22.13% of total
variation. Original descriptors with the highest correlation coefficients in the first axis are
agriculture land use (0.96), channel width (-0.91), water discharge (-0.83) and stream depth (-
0.82). The second axis main descriptor is channel surface (0.86).
The linear regression between k values as response variable and the first PCoA axis as
independent variable was not significant (p = 0.13). However, LG2 showed a higher
sedimentation level and when this site was removed, we could see a negative relation between
these parameters (r
2
= 0.5754, p = 0.029).
Table 3. Analysis of variance for the breakdown rate among heavy, moderate
and light agriculture areas.
Agriculture intensity Stream reach Breakdown rate (± SE) r
2
Low agriculture LG1 0.0113 (± 0.1377) 0.8756
LG 2 0.0040 (± 0.08681) 0.8659
Moderate agriculture LG 3 0.0133 (± 0.06388) 0.9837
LG 4 0.0096 (± 0.2348) 0.8057
LEN 1 0.0056 (± 0.08385) 0.944
LEN 2 0.0047 (± 0.1526) 0.8316
Heavy agriculture LEN 3 0.0025 (± 0.2071) 0.8876
LEN 4 0.0033 (± 0.1736) 0.9261
Df S.S. M.S. F value P
Treatments 2 4.0864 *10
-5
2.0432 *10
-5
1.3823 0.3328
Residuals 5 7.3905 *10
-5
1.4781 *10
-5
20
LG1
LG2
LG3
LG4
LEN4
LEN1
LEN2
LEN3
-0.3
-0.2
-0.1
0
0.1
0.2
0.3
-0.6 -0.4 -0.2 0 0.2 0.4 0.6
Figure 2. PCA for the environmental variables shown on table 1. Three groups are
distinguished: light agriculture (LG1 and LG2); moderate agriculture (LG3, LG4, LEN1
and LEN2); and, heavy agriculture (LEN3 and LEN4).
Chemical and physical water composition
Chemical analyses detected differences in five of the measured variables (Table 4).
Dissolved oxygen (DO) and pH-value were slower in high agriculture than in the other two
areas (p < 0.001). DO mean level was of 5.55 mg/L and pH-value was 6.85 at high agriculture
sites. In moderate and low agriculture mean pH level ranged from 7.25 to 7.40 and DO mean
concentration level ranged from 7.2 to 7.27 mg/L. Nitrate concentrations were higher in high
agriculture (1.45 mg/L) than in moderate (0.1 mg/L) and light agriculture (0.17 mg/L) areas (p
= 0.001). Phosphorous, however, presented a higher concentration on both light (0.065 mg/L)
and moderate (0.1 mg/L) agriculture, while high agriculture presented 0.03 mg/L (p = 0.02).
Temperature, although not statistically different, was slightly higher in the heavy agriculture
areas. Turbidity, on the other hand, showed a slight increase in the low agricultural areas.
Water discharge was more pronounced at both light and moderate agriculture areas (p = 0.01).
21
Table 4. Chemical and physical variables (mean values ± standard deviance) for light, moderate and heavy
agriculture areas. Small letters beside the values indicate significant difference (p < 0.05) between treatments.
Sediments
LG2, LEN3 and LEN4 presented a high amount of sedimentation, with silt and clay
texture presenting between 65 and 70% of the total sample (FEPAM, data unpublished). The
other sites had no more than 45% of silt and clay and a minimum of 20% of these two
components. High levels of sedimentation led to a negative effect on leaf breakdown (p =
0.043) (Table 5). Backwards stepwise multiple regression data showed that sand texture is the
best predictor of the model (R
2
= 0.7812, p = 0.002225; Table 6; Figure 3). Sand is negatively
related to silt and clay (Table 7).
Table 5. Analysis of variance between sediment accumulated sites.
Table 6. Sediment backward stepwise multiple regression analysis.
Leaf breakdown rates (as k values) are the responsible variables and
sediment texture (sand, gravel, clay and silt) are the independent
variables. Sediment original data (FEPAM, unpublished data) was used
after arcsine square root transformation. Only the last significant term
is shown (95% CI, n = 8).
Estimate Std. Error t value P
Intercept -0.006442 0.002683 -2.401 0.05320
Sand 0.019721 0.003868 5.098 0.00222
Variables Light agriculture Moderate agriculture Heavy agriculture p
Temperature (°C) 20.8 (± 2.6) 20.9 (± 1.8) 21.6 (± 2.8) 0.75
pH 7.2 (± 0.3)a 7.4 (± 0.1)a 6.8 (± 0.3)b <0.001
Dissolved O
2
(mg/l) 7.3 (± 0.7)a 7.2 (± 0.3)a 5.5 (± 0.7)b <0.001
Specific conductance (µS/cm) 63.5 (± 11.3) 62.5 (± 12.6) 63.2 (± 14.8) 0.97
Total Phosphorus 0.0 (± 0.0)a 0.1 (± 0.0)a 0.3 (± 0.0)b 0.02
Nitrate (NO
3
) 0.2 (± 0.2)a 0.1 (± 0.1)a 1.4 (± 1.4)b 0.001
Total Nitrogen
1.20 (± 0.52) 1.2 (± 0.4) 1.6 (± 1.3) 0.36
Turbidity 13.8 (± 5.05) 12.8 (± 9.9) 10.9 (±8.5) 0.58
Discharge (m
3
/s) 3.3 (± 2.8)a 2.1 (± 2.1)a 0.0 (± 0.0)b 0.01
Df Sum Sq Mean Sq F value P
Sediment 1 5.9502 *10
-5
5.9502 *10
-5
6.4598 0.04398
Residuals 6 5.5267 *10
-5
9.2110 *10
-6
22
Figure 3. Sediment backward stepwise multiple regression significant term. Independent
variables original data were used after arcsine square root transformation.
Table 7. Correlation between sediment components. In
bold are shown the components which presented
statistically significant correlations (p < 0.05).
Gravel Sand Silt Clay
Gravel 1
Sand 0.2091 1
Silt -0.60719 -0.86102 1
Clay -0.40288 -0.81568 0.68233 1
Discussion
In this work, we could not find a clear relationship between leaf breakdown and land
uses. However, we could make some speculations. First of all, the categories made after
landscape evaluation were not efficient as indicators of the ecological process. Royer and
Minshall (2003) attested that the use of leaf breakdown rates for biological assessments
should not be used for describing different scales than the reach scale where the leaves are
23
being exposed. Hawkins and Vinson (2000) also described the problem of weakness of land-
use classifications in bio-assessments to the invertebrates’ independent and continuous
variation over environmental gradients. As invertebrates are important feeders on leaves
(Kaushik & Hynes, 1971; Webster & Benfield, 1986), it is supposed that this non-detectable
variation when streams were a priori classified could have affected the non correspondence
between land use and breakdown rate. According to Whittier et al. (2007), even in choosing
reference sites large mistakes can be done, as reference sites not always follow researches
“expectance”.
The studied catchment presents an almost homogeneous chemical composition of
elements. Data collected by Rio Grande do Sul State Environmental Agency about metal
concentration on the substrate of the same sites studied here, detected only a small increase of
these compounds towards the river mouth but no real difference between sites (FEPAM,
unpublished data). Chemical composition differed for a few parameters from the heavy
agriculture to the moderate and light agriculture (Table 3). Although no significant difference
was observed for the breakdown rates, leaves in the heavy agricultural land-use type
presented a slower decomposition rate than the other agriculture areas (Table 2). Greater
amounts of phosphorous were detected in light and moderate agriculture in comparison to
heavy agriculture, which can have positively affected the leaf breakdown of these sites
(Suberkropp & Chauvet, 1995). Nitrate, on the other hand, was higher on heavy agriculture
than on moderate and light agriculture areas which attests the heavy agriculture sites to be
more influenced by agriculture itself and by live-stock grazing. However, several studies have
found a positive relation between nutrient enrichment (e.g., nitrogen and phosphorus) and leaf
breakdown rates (Robinson & Gessner, 2000; Ferreira et al., 2006; Bergfur et al., 2007).
Both low levels of dissolved oxygen and low pH-value presented on the heavy
agriculture areas can be explained by the slow discharge observed on these sites. These
physical-chemical properties can limit the macroinvertebrates activities, thus reducing their
abundance and richness, which could have influenced also shredders occurrence.
Another issue that influenced the non-observable relationship between litter
decomposition and environmental gradient is that one of the light agricultural areas (LG2)
presents a dam downstream the sampling site and a lack of riparian vegetation. These factors
contributed to a higher level of sedimentation which affected negatively the breakdown rates
(Table 5). Other works have also reported the negative affects of sedimentation to breakdown
24
rates (Niyogi et al., 2003, Sponseller & Benfield, 2001). For Sponseller and Benfield (2001),
sediments were the main cause of the non-detectable relation between breakdown rates and
land-uses. This negative relation found between sediments and breakdown rates in our work
was further examined as to verify which amount and which type of sediment interferes in the
decomposition rates. From the four sediment categories, sand was the only significant texture
related to breakdown rates and these two variables were positively related (Table 6; Figure 3).
Zweig and Rabeni (2001) found sand related to high sediment deposition and thus
responsible for decreasing invertebrates’ densities and richness. However, in their work sand
was related to silt and other small particles while our work detected higher presence of sand
related to lower amounts of silt and clay (Table 7) and, thus, lower sedimentation level.
Santmire and Leff (2007) also found larger particles related to higher microbiological
colonization and Sponseler and Benfield (2001) reported smaller particles as silt and clay to
be responsible for burying the leaves and slowing down its decomposition. From our data set
sand was an indirect (through negative relation to silt and clay) but significant describer (by
its positive relation with breakdown rates) of the effects of sedimentation on leaves
decomposition.
Our linear regression between PCA main axis and k values, after removing LG2,
showed a relation between some environmental variables (agriculture status, channel width,
water discharge and channel depth) and the breakdown rate. Although the relation detected
between PCA first axis and the k values without LG2 would support the idea that agriculture
status affects breakdown rates, most properties influencing this linear regression are related to
channel morphology and hydrology (channel width and depth and water discharge). Water
discharge, which can be affected by channel morphology, is known to affect breakdown rates
by physical abrasion (Niyogi et al. 2003) and this was the case of the low and moderate
agriculture areas. In the studied reaches, polluted areas are mainly located in upper sites of the
catchments, which present lower channel width and channel depth and thus, lower water
discharge. In these areas there are better soils available for growing vegetables and so there
are no upstream sites that present no human disturbance in this bio-region. This catchment’s
characteristic caused a problem with our data since pollution in this area is negatively related
to water discharge (FEPAM, unpublished data).
In the last few decades many works have focused on leaf breakdown in running water
systems due to its ecological importance especially to streams. As a consequence,
25
assumptions of its ecological process to the water bio-assessments have been constantly
reported (Gessner & Chauvet, 2002, Royer & Minshall, 2003). Some works have found a
relation between water pollution (e.g., metal contamination or organic input) and leaf
breakdown rates (Niyogi et al., 2003, Pascoal et al., 2003). However, some few authors trying
to find a relation between breakdown rates and land use gradient have verified that this
relation is to be at least discussible (Hagen et al. 2006, Sponseller & Benfield, 2001). Paul et
al. (2006) found fast breakdown rates in agricultural areas related to nutrient enrichment and
fast breakdown rates in urban areas related to storm runoff, while Hagen et al. (2006) found
low breakdown values in both heavy agriculture and forest sites. We found a relation between
sedimentation and leaf breakdown rates but not directly with agricultural areas, as destruction
of the riparian zone is not necessarily related to the agriculture practices.
Intensive agricultural fields may change an ecosystem structure through increasing of
nutrients (Lowrance et al., 1984; Smith, 1992) and contamination by insecticides and other
chemical products that may affect the invertebrates life-cycle (Niyogi et al., 2003, Gafner &
Robinson, 2007). On the other hand, the absence of a vegetated riparian zone leads to two
contradictory effects. First, it allows the increasing of algae biomass around leaves because of
more light availability, thus accelerating breakdown (e.g., Franken et al., 2005). In the same
way, the lack of a riparian zone will contribute for increasing the organic input from the
surrounding land uses (Smith, 1992; Watzin & McIntosh, 1999), when that is the case. As an
opposite effect, there is a higher sedimentation level at these areas which can bury leaves and
then slow down their breakdown rate (Sponseller & Benfield, 2001; Niyogi et al. 2003), or at
least, influence negatively invertebrates’ occurrence (Niyogi et al., 2007) as it was observed
in our work. For Niyogi et al. (2007) the restoration of the riparian zone could be an effective
improvement for stream habitats and invertebrates health.
Royer and Minshal (2003) argued that when leaf breakdown is to be used in biological
assessments, care should be taken as this ecological process is more directly related to the
reach scale than to catchment scale. Thus, we attest that caution should be taken for using this
ecological process in order to find relations between them and human disturbances. However,
as the studies have gone further, and as the problems on these studies so far have arisen, one
could start integrating them as to carefully understand such magnitude of impacts on this
important ecological process.
26
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30
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S
(Com Gilberto G. Rodrigues)
(Resumo aceito para apresentação oral no congresso Plant Litter Processing in Freshwaters 5,
Coimbra, Portugal, 2008, com possível publicação na International Review of Hydrobiology)
“Não é o mais forte da espécie que sobrevive,
nem o mais inteligente. É aquele indivíduo que
é mais adaptável à mudança”.
Charles Darwin
31
Abstract
Shredders are important components of litter decomposition on temperate regions.
However, there is a paucity of this feeding-group in warmer areas, especially in the neotropics.
Some reasons have been hypothesized for this scarcity of shredders, as (i) lower leaf quality,
(ii) hydrological factors influencing litter decomposition and (iii) high occurrence of predators
controlling shredders abundance in tropical and sub-tropical regions. This work aimed to
verify the presence of shredders on litter in a neotropical stream in southern Brazil and to
check to which extent decomposition rate is related to shredders occurrence. We placed litter-
bags in eight stream reaches and retrieved four samples each month, until six months of
exposure. Litter-bags were made of 4g of Ocotea puberula (Lauraceae) leaves collected
nearby the studied catchment. Reaches were chemically similar and were divided to its
discharge values, as our experiment presented a high water discharge and a low water
discharge treatment. Litter breakdown rate was significantly different between treatments (p =
0.03). Macroinvertebrates assemblage values, as total number, density and taxa richness did
not differ between treatments (p > 0.05) but abundance and density of shredders differed
significantly (p = 0.01; p = 0.04, respectively). The highest discharge reaches presented the
fastest decomposition but the lowest number of shredders. On the other hand, slow discharge
reaches presented low decomposition rates and greater amount of shredders. The rate between
predators and other invertebrates was similar between treatments (p = 0.37) suggesting that
predators were not important in controlling shredders abundance. These findings suggest that
shredders are not so important for leaf decomposition in the neotropical region and, in
opposition to what happens in temperate areas, shredders need resource availability (which is
promoted by lower physical breakdown rates).
Key-words: leaf decomposition, neotropics, shredders, resource, water discharge.
32
Introduction
Studies conducted in the last thirty years in temperate streams have verified the
importance of shredders to the detritus chain on aquatic systems (e.g. Kaushik & Hynes, 1973;
Fisher & Likens, 1973; Petersen & Cummins, 1974; Vannote et al., 1980). This ecological
food-chain starts on litter input from surrounding areas on streams (Webster & Benfield, 1986)
and further colonization of the leaves by fungi hyphomycetes (Gessner et al., 2007).
Following this first step, shredders colonize leaves and feed on these fungi and/or in the
leaves’ mesophyllo (Graça, 2001; Wantzen & Wagner, 2006).
Although fungi are a very important step on leaf decomposition (Gessner et al., 2007),
shredders are responsible for most of the leaves’ breakdown (e.g., Niyogi et al., 2001).
Nevertheless, on this detritivorous food-chain, shredders main function is to available the
great amount of resource that enters temperate streams for a large variety of other organisms,
once the coarse particulate matter will be reduced to fine particulate matter (Graça et al.,
2001). Thus, collectors and filters located on lower areas of the streams will be directly linked
to this step (Allemano et al., 2007).
It is claimed that shredders are either scarce in tropical regions (Wantzen & Wagner,
2006), especially on neotropic areas, or they play a minor role on leaf decomposition
(Mathuriau & Chauvet, 2002; Moretti et al., 2007). Contradictorily, there is a great input of
leaves into these systems during all year, which one could expect to be a reason for the
presence of shredders. Some authors suggested that leaves in the tropics have more
unpalatable substances, as secondary compounds (Lavelle et al., 1993; Aerts, 1997), which
would inhibit the feeding of aquatic invertebrates on them. Others, such as Wantzen &
Wagner (2006) argue that tropical regions present a greater amount of predators and so,
invertebrates in general should have a more diversified feeding diet.
Hydrological factors would also be reasons for the paucity of shredders (Winterbourn
et al., 1981; Youle, 1996) as tropical rain storms can easily “wash” stream substratum, taking
the food source away from the invertebrates, and again, inhibiting its feeding behavior. On
one hand, some authors have suggested that reduced discharge should reduce invertebrates
density and richness (e.g., Dewson et al., 2007b) or invertebrates abundance (e.g., Wood et al.
2000) by altering availability and suitability of instream habitats (Dewson et al., 2007a). On
33
the other hand, an increasing in invertebrates abundance due to slower current velocity was
also detected by Suren et al. (2003). In this way, hydrological effects on macroinvertebrates
assemblage is still discussible (see Dewson 2007a for a review). We hypothesize that, besides
leaf quality, shredders abundance will be related to resource availability mediated by
hydrological factors if predators are not controlling their abundance.
Methods
Sites characterization
The stream used in this experiment is Lajeado Grande stream. It is located in southern
Brazil and presents a sub-tropical climate (27ºS, 54ºW). Rain fall average is of 1800 mm a
year and mean temperature in summer and winter are 25 °C and 8 °C, respectively. This
catchment is on deciduous broadleaf forest area which originally covered great part of
Southern and Central Brazil. Leaves fall during all year and Lauraceae and Myrtaceae
comprise the main tree families presented on this region.
In order to verify water discharge influence on shredders occurrence, we selected six
sites differing in discharge values (Table 1). They were selected in the same basin to avoid
differences in water chemistry (Table 2) but located as far from each other as possible, to have
them as independent sites. After the beginning of the experiment, one of the slow discharge
sites had to be removed from analyses as higher sedimentation was observed which changed
its characteristics making it difficult to compare with the other five sites. All chemical and
physical data was taken by Rio Grande do Sul State Environmental Agency (FEPAM).
Table 1. Water discharge (m
3
/s) measured
monthly between October 2004 and January
2005. Data was kindly supported by FEPAM.
Treatments Water discharge in m
3
/s (± sd)
1 0.97 (± 0.57)
2 3.28 (± 2.87)
34
Table 2. Mean chemical variables (± sd) taken monthly from October
2004 to January 2005 on each sampling site (Treatment 1: n = 3;
treatment 2: n = 2). Data was kindly supported by FEPAM.
Treatment 1
average (± s.d.)
Treatment 2
average (± s.d.)
p
Temperature 20.9 (± 2.6) 20.5 (± 1.73) 0.7418
pH 7.3 (± 0.3) 7.3 (± 0.11) 0.9034
Dissolved Oxygen 7.3 (± 0.8) 7.3 (± 0.35) 0.9007
Conductivity 63.6 (± 11.3) 56.5 (± 9.34) 0.1621
Total phosphorus 0.1 (± 0.0) 0.1 (± 0.03) 0.3557
Total nitrogen 1.2 (± 0.5) 1.2 (± 0.36) 0.954
DBO 2.2 (± 0.3) 2.1 (± 0.14) 0.1717
Turbidity 13.9 (± 5.1) 13.6 (± 4.8) 0.9177
Total solids 106.8 (± 42.8) 109.0 (± 32.6) 0.9012
Leaf-bags
Leaf-bags were made of 4g (± 0.01g) of Ocotea puberula (Lauraceae) leaves collected
just after abscission and air-dried until necessary. Bags were 30 x 20cm large and presented a
mesh size of 10mm to allow the presence of invertebrates. As some studies have shown that
shredders in the neotropics are related to wetland pools or to river banks (Wantzen & Wagner,
2006; Ríncon & Martínez, 2006; Graça et al., 2001) our leaf-bags were intentionally placed
next to the stream banks, divided in four riffles which were considered as replicates. At each
sampling date (i.e., after 15, 30, 60 and 90 days from incubation) one leaf-bag from each
riffle was retrieved and stored in a cooler for transportation to the laboratory. Then, leaves
were gently washed over a 200µm sieve to remove invertebrates for further analyses.
Invertebrates were kept in 70% alcohol and identified using stereomicroscope. Leaves were
oven-dried for 48 hours and weighed to the nearest 0.01g. Breakdown rates were then
calculated on a negative exponential formula (Petersen & Cummins, 1974).
Functional types: Shredders and predators
Macroinvertebrates were identified to genera level according to Fernandéz and
Dominguéz (2001) for most insects and according to Costa et al. (2004) and Pes et al., (2005)
for most of the Odonata and Trichoptera, respectively. Plecoptera was identified following
Olifiers et al. (2004). Total genera richness, total abundance and density were measured and a
ratio between predators and other macroinvertebrates number was also calculated.
35
Statistical analysis
Breakdown rates obtained from O. puberula leaves in slow and fast current velocity
sites were compared in a one-way analysis of variance (1-way ANOVA). Analyses of Co-
Variance (ANCOVA) was used for comparing genera richness, macroinvertebrates
abundance and density between treatments where time was used as covariate. Predator and
prey ratio between treatments and during exposure time was analyzed on a two-way ANOVA.
Shredders abundance and density were compared between treatments and time (used as a
covariate) on an ANCOVA analysis. All statistical analyses were conducted on R program
(2006).
Results
Decomposition
Leaf breakdown differed between treatments (p = 0.004) and during time (p < 0.001).
At the fast discharge, O. puberula breakdown was two times faster than at the slow discharge
sites. At the fast current sites average k was of 0.0114 and in the slow current sites average k
was of 0.0053 (Table 3).
Table 3. Breakdown rates (log-transformed) obtained from Ocotea
puberula leaves in slow and fast current velocity sites. (p = 0.004
between treatments and p < 0.001 during time)
Macroinvertebrates assemblage
There was a general trend for invertebrates presenting lower values at the end of the
experiment. However, genera richness, density and abundance did not differ between
treatments (p > 0.05) nor during time (p > 0.05) (Table 4).
The ratio between predators and other invertebrates did not differ between treatments
(Table 5) and was around 1:8 (predators:invertebrates). Most predators belong to Odonata
Current velocity Site Breakdown rate (± SE) r
2
Fast discharge F1 0.0113 (± 0.1377) 0.8756
F2 0.0133 (± 0.06388) 0.9837
F3 0.0096 (± 0.2348) 0.8057
Slow discharge S1 0.0056 (± 0.08385) 0.944
S2 0.0047 (± 0.1526) 0.8316
36
group, from both Zygoptera and Anisoptera sub-orders which main families were
Calopterygidae, Coenagrionidae and Libellulidae. Other predators found were Hydroptilidae
Neotrichia sp. and Hydroptila sp. and Perlidae Anacroneuria sp. Invertebrates and predators
ratio did not differ during exposure time.
Table 4. Analyses of Co-Variance (ANCOVA) for genera richness, macroinvertebrates abundance and
density. Numbers in genera richness are related to total number of genera found. Abundance is related to
total macroinvertebrates number found in litter-bags. Predictor variables are treatment and time (covariate).
Response variable Independent variable D.f. Sum Sq Mean Sq F value P
Treatment 1 1 1 0.0002 0.9890
Time 3 10464 3488 0.9256 0.4524
Treatment:Time 3 2745 915 0.2428 0.8651
Abundance
Residuals 15 56526 3768
Treatment 1 36663 36663 3.0674 0.1003
Time 3 86113 28704 2.4015 0.1083
Treatment:Time 3 46961 15654 1.3097 0.3079
Density
Residuals 15 179287 11952
Treatment 1 15.653 15.653 3.5962 0.07734
Time 3 7.816 2.605 0.5986 0.62575
Treatment:Time 3 13.806 4.602 1.0573 0.39641
Genera richness
Residuals 15 65.292 4.353
Table 5. Analysis of variance (ANOVA) of predator and prey ratio between
treatments and during exposure time.
Predator:prey ratio Df Sum Sq Mean Sq F value p
Treatment 1 0.000895 0.000895 0.803 0.3746
Time 3 0.005641 0.001880 1.6876 0.1821
Residuals 48 0.053480 0.001114
Shredders found on the leaf-bags were mainly insects but some other groups were
considered shredders as well. Crustaceans found in this work (Aegla sp. and Trichodacthylus
sp.) belong to Decapoda group and are usually classified as macroconsumers that can feed on
invertebrates and act as shredders, especially when searching for miners into the leaves.
Besides these, true shredders as Phylloicus sp. and Limnoperla sp. from Calamoceratidae and
Grypopterygdae families, respectively, and dipterans Tipulidae were found on the leaf packs.
Shredders abundance and density, however, differed significantly between treatments
(p < 0.05) but not during time (p > 0.05) (Table 6; Figure 1 and 2). Most shredders were
found on the slow current sites and belong to Grypopterygidae (Plecoptera), Calamoceratidae
(Trichoptera), Tipulidae (Diptera), Trichodactylidae and Aeglidae (Crustacea). The main
difference between treatments is that Phylloicus sp. (Calamoceratidae) was present at a higher
grade in the slow stream flow sites and Trichodactylidae (Crustacea) was only found in these
slow current sites (Table 7).
37
Figure 1. Analysis of covariance (ANCOVA) showing higher shredder
density in the low discharge sites than in the high discharge sites. Density
does not change during time. (Treatments: p = 0.0462; Time: p = 0.3268).
Figure 2. Analysis of covariance (ANCOVA) showing higher shredder
abundance in the low discharge sites than in the high discharge sites.
Abundance does not change during time. (Treatments: p = 0.01562; Time: p =
0.078).
38
Table 6. Analysis of covariance among Shredders mean number, treatments and time (as
covariate). In bold, the statistically significant terms (p < 0.05).
Response
variable
Independent
variables
D.f. Sum Sq Mean Sq F value p
Time 3 21.568 7.189 1.1811 0.3268
Velocity 1 25.492 25.492 4.1882 0.0462
Shredder
density
Residuals 48 292.162 6.087
Time 3 73.61 24.54 2.4091 0.07852
Velocity 1 64.00 64.00 6.2839 0.01562
Shredders
abundance
Residuals 48 488.85 10.18
Table 7. Shredders richness (X indicates presence and O indicates absence of a certain taxa. Fast discharge
treatment: n = 3; slow discharge treatment 2: n = 2)
Trichoptera Plecoptera Diptera Crustacea
Calamoceratidae
Gripopterygidae
Tipulidae
Aeglidae Trichodactylidae
Time Treatment
Phylloicus sp. Limnoperla sp.
Sp1 Aegla sp.
Trichodactylus sp.
Fast X X X 0 0
15
Slow 0 X X 0 X
Fast X X X X 0
30
Slow X X X X X
Fast 0 X X 0 0
60
Slow X 0 X 0 X
Fast 0 0 0 0 0
90
Slow X 0 0 0 X
Discussion
Functional role: Shredders, predators and resource quality
There is a non-agreement of the current literature on the feeding-functional
classifications for some taxa found on this work. Smicridea sp. can be considered both as
collector-gather (Poff et al., 2006) or as generalist (Wantzen & Wagner, 2006); Trichorythods
sp. could be classified as collector-gatherer (Poff et al., 2006) or as collector that could
behave as shredder or scraper (Wantzen & Wagner, 2006); Nectopsyche sp. is considered as
generalist herbivore by Poff et al. (2006) although some Leptoceridae are considered
shredders (e.g., Graça et al., 2006; Wantzen & Wagner, 2006); and Pyralidae presents the
same problem as Nectopsyche sp. besides the fact that Pyralidae has its classification based on
family level. So, these taxa were not considered as shredders even though the number of
Nectopsyche sp., in especial, found in this work would make the differences between
treatments even larger, helping our results to become more easily confirmed (data not shown).
39
Several reasons have been hypothesized in relation to the paucity of shredders in the
neotropics in comparison to temperate regions: (i) lower quality of food, regarding to leaf
toughness and presence of tannins and secondary compounds (Graça & Barlocher, 1998;
Wantzen et al., 2002; Poorter et al., 2004); (ii) higher amount of predators; and, (iii)
hydrological factors that could disturb aquatic systems in higher frequency (as in tropical rain
storms) (Wantzen & Wagner, 2006). These factors linked with the non-correspondence of
larvae developing time with resource input, as it happens in temperate regions, could make
shredder-specialist behavior even more difficult to exist.
Even though leaves in tropic regions can have more secondary compounds, many
organisms can avoid its chemicals as it has been shown for temperate invertebrates (Canhoto
& Graça, 1999) or they can even choose among leaf species (Rincon & Martínez, 2006).
Rincon and Martínez (2006) showed that Phylloicus sp. could select leaves among native
species which differed in nutrient quality, in Venezuela. Graça et al. (2001) showed that
tropical shredders survived and grew in similar ways on leaves from both temperate (Alnus
glutinosa (L.) Gaertn.) and tropical regions (Hura crepitans L.). In a tropical stream Ardón
and Pringle (2008) demonstrated that leaf secondary compounds did not influence leaf
breakdown in Costa Rica, but structural components as cellulose did.
We used the same leaf species in all reaches and so leaf quality should not influence
differences between treatments. The very low decomposition rate verified on this work (k =
0.0047 and 0.0052), which are included among the smallest rates already verified in the
neotropics (Rueda-Delgado et al., 2006), is in agreement with the decomposition rate of
another Ocotea used in an experiment in Brazilian Cerrado (Moretti et al., 2007). Moretti et
al. (2007) found, for this species, the highest density and biomass of invertebrates, although
in Cerrado streams a decomposition rate of 0.005 was considered as fast.
It should be expected, though, that either predaceous or lack of resource would
account for treatments exhibiting a poor shredder abundance. In this work we could verify
that predators’ abundance was similar in both treatments (p > 0.1) and the predators:prey ratio
did not present variation between treatments (p > 0.05).
Thus, the last reason for influencing shredders abundance would be the lack of
resource. It has been shown that neotropic shredders are abundant in lateral wetland pools
(Wantzen & Junk, 2000) or in slow flowing areas (Wantzen et al., 2002; Cheshire et al.,
2005). We sampled leaf-bags next to the stream bank where, following this assumptions,
40
shredders would be more abundant, and we avoided middle parts of the streams. With the
same chemical composition, the reaches studied here differed only to its water flow
(discharge). As macroinvertebrates abundance and density were similar between treatments (p
> 0.05 for both parameters) but shredders abundance and density were higher in the slow
discharge treatment (p = 0.03) we could assume that leaves presented higher breakdown due
to physical fragmentation (disruption) and not to invertebrates feeding (Figure 2 and 3).
Although most literature agree that flow reductions (artificially or natural) reduce
invertebrates richness because of reducing habitat availability (Dewson et al., 2007a), the
slow flow detected here was only enough to reduce physical fragmentation of leaves and not
to change in-stream habitats (e.g., reducing wetted areas).
Rueda-Delgado et al. (2006) verified that leaves of different qualities in a headwater
Amazon stream were faster decomposed at the fast-discharge season, where Amazon River is
low and stream discharge is more variable than the high-water period, when stream discharge
is very low. Although the proportion of functional feeding groups found on their work did not
differ statistically between rain seasons, the proportions of shredders as Phylloicus was
slightly higher in the slow discharge season than in the fast discharge season. However, a
work on Amazonian litter-banks during slow-flowing season revealed a high diversity of
invertebrates on litter but no litter-feeding specialist was found (Henderson & Walker, 1986)
as most organisms were shrimps and the primary consumers fed mainly on algae or fungi on
litter, but not on litter itself.
Cheshire et al. (2005) verified a higher richness and biomass of shredders in pool
habitats rather than riffles in Australia. Higher occurrence of shredders in slow flowing
streams, in South America, was verified by Chara et al. (2007), who detected a high
importance of shredders on leaves in a slow flowing stream in Colombia, where this
functional group represented over 50% of invertebrates biomass on leaf bags.
Some studies on other tropical regions have also verified a paucity of shredders in
streams, as in Kenya (Dobson et al., 2002) and in Papua New Guinea (Yule, 1996). Besides
Australian tropical streams (Cheshire et al., 2005), most reports indicate a climate
correspondence effect for this feeding-group around the world. Winterbourn et al., (1981) had
already noted a paucity of shredders in New Zealand streams for which they pointed New
Zealand non-retentive instable stream environments as an explanation for this effect. Hence,
according to provided literature for the tropics, in special to that data from the neotropics, it
41
seems that in this region there is a change on the current overview about shredders influence
on leaves (Graça et al., 2001), as shredders do not play a major role on decomposition
(Mathuriau & Chauvet, 2002) but may be influenced by resource availability. However, we
emphasize that more studies are needed for neotropical regions to verify to which extent
shredders might influence or not litter decomposition and to understand which shifts might
occur in shredders distribution among habitats (Cheshire et al., 2005) or from slow
decomposing to fast decomposing sites.
One ecological consequence outcome from this different shredder-resource-water flux
link is their possible occurrence fragility in the neotropics. This is because in spite of several
measures that have been undertaken in the last few years aiming conservation and
restauration, the neotropical region are still often subjected to non-sustainable development
(e.g., Malhi et al., 2008). This fact encourages land exploitation for agriculture and
destruction of the riparian vegetation by forest logging which can increase water flux events
and thus move out the resource needed for shredders (Naiman & Décamps, 1997).
Kreutzweiser et al. (2008), for instance, could detect effects of logging on macroinvertebrates
assembly and on leaf-decomposition even under best management practices in Canada. In this
scenario, the few shredders that occur in these tropical and sub-tropical regions can become
even more endangered, which can lead to the break of a very important trophic link, even
though small in the neotropics (Graça et al., 2001), between headwater shredders and lower
stream reach invertebrates.
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46
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(Com Gilberto G. Rodrigues)
(Artigo em preparação para ser submetido à Acta Limnologica Brasiliensia)
“O rio da minha aldeia não faz pensar em nada.
Quem está ao pé dele está só ao pé dele.”
Poemas de Alberto Caeiro
Fernando Pessoa
47
Abstract
The use of macroinvertebrates in biological assessments has been widely suggested by
researchers and environmental agencies. While some countries have it well developed, as
England, USA and Australia, other ones, especially in southern hemisphere, still do not have
these metrics established. In Brazil, studies focusing macroinvertebrates as biological
indicators have just started and, due to Brazilian continental size, these studies are valid for a
few areas of the country. In this work we assessed eight reaches in southern Brazil using
biological indexes (BMWP and ASPT) already proposed from Brazilian Cerrado and
Argentinean Patagonia. Our studied catchment presented low index values for both Cerrado
and Patagonean scores. However, differences were found between both indexes. One site was
classified as more impacted when using Argentinean values but was classified as less
impacted when used Cerrado values. This first approach of bio-assessment in southern Brazil
shows the urgent need of more ecological and taxonomical studies so that one could classify
more accurately running waters in this region of Brazil.
Key-words: Biological assessment, BMWP, ASPT, macroinvertebrates, subtropical
climate, southern Brazil.
48
Introduction
Benthic macroinvertebrates are very useful organisms in environmental assessments.
They are easy to be collected and handled, have appropriate identification keys, are abundant
and present many different ecological traits (Rosemberg & Resh, 1993). Because of this, the
use of macroinvertebrates in biological assessments has been well developed, especially in
northern hemisphere countries (Rosenberg & Resh, 1993) as several indices have been
proposed (Stribling et al., 2008). Southern hemisphere, and South America in especial, in the
other hand, is just starting to have these indexes used (Cota et al., 2002; Junqueira & Campos,
1998; Junqueira et al., 2000, Miserendino & Pizzolón, 1999) although countries as Australia
and New Zealand can be considered as outliers.
These bio-assessments follow well-studied protocols which classify the waters into
classes after the observation of specific benthic invertebrates taxa occurrence. Generally,
these protocols are based on family-taxa level, due especially to its easier identification when
compared to genera and species level (Barbour et al., 1996; Chessman et al., 2007). To date
most researches and protocol developing have been conducted based on the Biological
Monitoring Working Party index (BMWP) developed in England (Armitage et al., 1983).
This index scores macroinvertebrates taxa from one to ten, as the lowest values stand for poor
water quality indicator organisms and the highest values stand for good water quality
indicator organisms.
Another index used is called ASPT (Average Score per Taxon) and is calculated based
on the values obtained from BMWP but divided by the number of taxa found in each site.
Some assessment analyses have also used the EPT values, which stands for three main groups
of macroinvertebrates: Ephemeroptera, Plecoptera and Trichoptera (e.g., Zweigg & Rabeni,
2001). However, the use of an index at the order level from only three orders is less sensitive
in water analyses once these orders can have many families that respond to different levels of
pollution (but see Melo, 2005).
In this work, we assessed water quality of several stream reaches in southern Brazil,
through BMWP modified values from Cerrado and Argentinean Patagonia. Our goal is to
identify which index is more appropriated for this region and to provide literature with new
data from macroinvertebrates biological assessments.
49
Material and Methods
Study site
The studied catchment is located in southern Brazil, in a sub-tropical area influenced
by the Deciduous Forest. The stream, called Lajeado Grande, presents 83.4 km length and
mean depth of 0.6 m. Its basin covers 525.38 km
2
, presenting its headwaters on southern
Brazil Plato and its mouth in Uruguay River. The area presents an intense agricultural activity,
most of which uses the riparian zone for cropping. The most frequent tree species on this area
belong to Lauraceae and Myrtaceae families.
Eight reaches (Table 1, Figure 1) were chosen to have the aquatic macroinvertebrates
sampled. These sites present different land uses around the stream, including pig growing
farms and soy/corn monocultures. Water chemistry was measured monthly and substrate
composition was verified in October 2004 by Rio Grande do Sul State Environmental Agency
(FEPAM). The chemical and physical measurements were done for pH-value, dissolved
oxygen, electrical conductivity, turbidity, temperature, water discharge, total phosphorus, total
nitrogen, nitrate. Substrate was classified into percentages of gravel, sand, silt and clay
(FEPAM , unpublished data).
Macroinvertebrates
For sampling the invertebrates we used litter-bags attached to the stream bank. This
procedure should allow us to collect both invertebrates that use leaves as a substratum or
those invertebrates which actively feed on leaves. The purpose on doing so is to be able to
analyse invertebrates specifically related to the very important feeding group of shredders
which, for feeding on leaves, can be good indicators of human activities on the riparian zone.
The 10mm mesh-size litter-bags were made of 4g of Ocotea puberula leaves. These
leaves were collected from one single tree next to Lajeado Grande catchment. Bags were
placed on 6 reaches of the stream. In each reach, groups of ten litter-bags were divided into
four points on the stream bank where riffles were present. These points were used as
replicates in statistical calculations. At 15, 30, 60 and 90 days the 4 litter-bags from each
reach were retrieved from water into plastic bags and stored in a cooler for transportation to
the lab. Then, leaves were gently washed with tap water over a 250µm sieve and the material
50
was kept in 70% alcohol for further invertebrates’ identification. Fernandéz and Dominguéz
(2001) was the main source for identifying Insecta. Costa et al. (2004) was used for most
Odonata, Pes et al., (2005) for Trichoptera and Olifiers et al. (2004) for Plecoptera. Mollusca
were identified by specialists.
Bio-assessment
For assessing the invertebrates assemblage we used literature source from Argentinean
Patagonia and from Brazilian Cerrado. Miserendino & Pizzolón (1999) were the reference
from Argentina and Cota et al. (2002) and Junqueira et al. (2000) were the reference from
Brazil (Table 1).
Statistical analyses
An one-way Analysis of Variance (1-way ANOVA) was used to compare both BMWP
and ASPT values among sampling sites. Whenever differences were significant, a posteriori
TukeyHSD test was then executed in order to verify which sampling sites differed. Time was
used as a block. The eighteen most abundant genera were related to the chemical and physical
measurements and to substrate composition in a Canonical Components Analysis (bi-plot cca).
51
Figure 1. Lajeado Grande catchment, southern Brazil. Sampling sites for aquatic macroinvertebrates analyses
are shown.
Results
BMWP and ASPT
From all taxa found on the Lajeado Grande sites, Noteridae, Leptohyphidae,
Megapodagrionidae, Protoneuridae, Hemiptera, Trichodactylidae, Hirudinea, Ampullariidae,
Nematyelminthe could not be included in any classification due to lack of taxonomic
information for this taxa. However, 30 taxa found on Lajeado Grande catchment were
included in the analyses. Both Cerrado and Patagonian indexes did not have values for four
taxa (Table 2).
On the BMWP values (Figure 2), differences among sites could be found only through
Patagonian index (Table 3). However, Tukey HSD a posteriori test did not find any
significant difference between sites. Time of sampling (related to litter decomposition) did not
differ for any of the indexes used.
52
Table 2. BMWP values from two distinct regions: Brazilian Cerrado (after
Junqueira et al., 2001 and Cota et al., 2000) and Argentinean Patagonia (after
Miserendino & Pizzolón, 1999). Some of the families were not present in one or
both the indexes shown below.
Class Order
Family Patagonia Cerrado
Insecta Trichoptera
Calamoceratidae 8 -
Hydropsychidae 5 6
Hydroptilidae - 7
Leptoceridae 10 7
Policentropodidae 7 7
Gyrinidae 3 5
Coleoptera
Elmidae 5 5
Dytiscidae 3 4
Ephemeroptera
Baetidae 6 5
Caenidae 4 4
Leptophlebiidae 10 10
Plecoptera
Gripopterygidae 10 10
Perlidae 10 8
Zygoptera
Calopterygidae - 8
Coenagrionidae 6 7
Anisoptera
Aeshnidae 6 8
Cordulidae - -
Gomphidae 7 5
Libellulidae 6 8
Lepidoptera
Pyralidae - 8
Diptera
Ceratopogonidae 4 4
Chironomidae 2 2
Simulidae 5 5
Tipulidae 5 5
Crustacea Decapoda
Aeglidae 5 -
Annelida Oligochaeta
Oligochaeta 1 1
Mollusca Gastropoda
Ancylidae 5 6
Hydrobiidae 3 -
Planorbidae 3 3
Bivalvia
Sphaeriidae 3 3
Table 3. Analysis of Variance (ANOVA) for the BMWP values from Argentinean Patagonia and
Brazilian Cerrado. Significant results are shown in bold, (p < 0.05).
BMWP Df Sum Sq Mean Sq F value P
Time 3 183.8 61.3 0.4936 0.688159
Site 5 3508.6 701.7 5.6545 0.000274 Patagonia
residuals 56 6949.6 124.1
Time 3 119.0 39.7 0.2096 0.8893
Site 5 2138.8 427.8 2.2609 0.0607 Cerrado
residuals 56 10594.8 189.2
For the ASPT values (Figure 3), both Cerrado and Patagonia indexes showed
differences among southern Brazilian sites (Table 4). As for the BMWP values, time did not
present difference for biological values in any of the indexes used. Through Tukey HSD a
posteriori test we could verify differences between LG 79 and LG 37 (p=0.0456869) from
53
Cerrado ASPT values and between LG 79 and LEN 7 (p = 0.0078286) from Patagonia ASPT
values.
Table 4. Analysis of Variance (ANOVA) for the ASPT values from Argentinean Patagonia and
Brazilian Cerrado. In bold, the significant results (p < 0.05).
ASPT Df Sum Sq Mean Sq F value P
Time 3 3.6475 1.2158 2.5812 0.06246
Site 5 7.2503 1.4501 3.0784 0.01592 Patagonia
Residuals 56 26.3781 0.4710
Time 3 0.7075 0.2358 0.4209 0.73870
Site 5 6.8641 1.3728 2.4502 0.04456
Cerrado
Residuals 55 31.3766 0.5603
Relation to physical and chemical variables
The CCA analysis for the eighteenth most important invertebrate families and the
chemical and physical variables of the streams showed significance for the permutated model
(Table 5). Some families were closed related to presence of silt/clay, as Tipulidae and
Gripopterygidae. Simulidae was related to high levels of dissolved oxygen, high turbidity and
presence of gravels. Oligochaeta and Leptohyphidae were related total nitrogen
concentrations, phosphorus concentrations and to sand texture. Elmidae, Ceratopogonidae and
Leptoceridae were related to high levels of phosphorus concentrations. Coenagrionidae,
Gyrinidae and Calamoceratidae were related to nitrate concentrations, high pH-value and silt.
Baetidae, Chironomidae and Caenidae were only related to temperature. Leptophlebidae was
related to water discharge and Sphaeridae was related to electric conductivity (Figure 4).
Table 5. Permutation test for CCA under direct model
Df Chisq F N. Perm P
Model 13 0.7302
1.8964
800 0.02875
Residual 4 0.1185
54
Figure 2. ASPT values for sampling sites from (a) Cerrado
and (b) Patagonia. Cerrado and Patagonian ASPT values
showed significant difference among southern Brazilian sites
(p < 0.05).
55
Figure 3. BMWP values for sampling sites from (a) Cerrado
and (b) Patagonia. Only Patagonian BMWP values showed
significant difference among southern Brazilian sites (p <
0.05).
56
LEN7 presented the highest score from Argentinean ASPT and BMWP values. This
site presented a greater number of Leptoceridae and Gripopterygidae organisms which score
ten on Argentinean BMWP index. These organisms are manly shredders and can indicate the
presence of riparian vegetation around the stream.
LG37, on the other hand, presented the lowest value from Cerrado ASPT and BMWP
index. This site presented the highest number of Oligochaeta and Chironomidae which are
indicators of poor water quality presenting the lowest values on Cerrado BMWP. LG9
differed from both mentioned sites and presented a large number of organisms, mainly
Coleoptera and Zygoptera.
Figure 4. CCA obtained for invertebrates` families and chemical and physical variables (arrows).
Invertebrate families: “Coena” = Coenagrionidae; “Calamo” = Calamoceratidae; “Hydropsy” =
Hydropsychidae; “Hydropt” = Hydroptilidae; “Leptoce” = Leptoceridae; “Elmidae”; “Gyrin” = Gyrinidae;
“Leptophl” = Leptophlebiidae; “Caenidae”; “Leptohy” = Leptohyphidae; “Baetidae”; “Gripopter” =
Gripopterygidae; “Coenagri” = Coenagrionidae; “Chirono” = Chironomidae; “Simul” = Simulidae;
“Cerato” = Ceratopogonidae; “Rhaghi” = Rhagionidae; “Oligoch” = Oligochaeta; “Sphaer” = Sphaeridae.
Chemical and physical variables: “temp” = temperature; “pH”; “DO” = dissolved oxygen; “cond” =
conductivity; “total_P” = total Phosphorus; “nitrate”; “total_N” = total Nitrogen; “turb” = turbidity;
“disch” = discharge; “gravel”; “sand”; “silt”; “clay”.
57
Discussion
Southern Brazilian sites presented low classification values from both Argentinean and
Brazilian indexes. The highest BMWP values were found on those sites where more shredders
were present. This feeding-group is related to areas that have well-preserved vegetation cover
and great amounts of litter input. According to Alba-Tecedor (1996), however, all sites should
be classified as contaminated water, ranking from 36 – 60 points on the BMWP index on both
Patagonian and Cerrado values.
Different classifications were obtained from Cerrado and Patagonian values to the
fauna found on Lajeado Grande stream. From the Cerrado index, only ASPT showed
differences among sites (p = 0.044). From the Patagonian values, however, both ASPT and
BMWP presented differences (p < 0.05) among sites.
All taxa were included, even rare taxa. Melo (2005) attested that the use of simplified
data-methods are mostly reliable when comparing streams at local scale, which was the case
in our study. Sickle et al. (2007), on the other hand, argues that the exclusion of rare taxa
affects results on bio-assessments. We also did not have reference sites as this catchment
presents a high disturbance over all region, and, in this case, reference sites could act as weak-
references (Whittier et al., 2007).
Through the CCA analysis we were unable to make a precise correspondence between
the physical and chemical variables and the invertebrate families, according to the BMWP
indexes obtained from South America. Elmidae and Ceratopogonidae families, for instance,
which are considered as indicators of low-intermediate water quality (ranging 4 and 5,
respectively, in the BMWP values of both Cerrado and Patagonia) were related to high total
phosphorus concentrations but Leptoceridae, which is considered as indicator of intermediate-
high water quality (scoring 10 and 7 in Cerrado and Patagonia indexes, respectively) was
related to high total phosphorus as well. In the same way, Gyrinidae, Coenagrionidae and
Calamoceratidae were related to high concentration of nitrate, high pH-value and presence of
silt texture, even though some can be considered as low water quality indicators (Gyrinidae)
and other as intermediate to high water quality indicators (Coenagrionidae and
Calamoceratidae). Baetidae, Chironomidae and Caenidae scored between 2 to 6 in Cerrado
58
and Patagonian values and were related to higher temperature which can occur in areas
without sufficient canopy cover, which allows sunlight to increase water temperature.
Although literature on biological assessment subject is very rich for Northern
hemisphere countries (see Rosemberg & Resh, 1993) we attest that more studies should be
conducted in South America in general, but in southern Brazil in especial. Once we have a
data-base of invertebrates distribution and their relation to water parameters we will be able to
standardize this method for biological assessments and make use of these organisms as water-
quality indicators. For now, the use of aquatic macroinvertebrates on biological assessments
may produce weak classifications for streams in regions where these organisms occurrence
and environmental tolerance is not yet well known.
References
Alba-Tecedor, J. 1996. Macroinvertebrados acuaticos y calidad de las aguas de los rios. In: IV
Simposio del agua en Andalucia (SIAGA), Almeria, 2:203-213.
Armitage, P. D., Moss, D., Wright, J. F. & Furse, M. T. 1983. The performance of a new
biological water quality score system based on macroinvertebrates over a wide range of
unpolluted running-water sites. Water Resource, 17:333-347.
Barbour, M. T., Gerristen, J., Griffith, G. E., Frydenborg, R., Mccarron, E., White, J. S. &
Bastian, M. L. 1996. A framework for biological criteria for Florida streams using benthic
macroinvertebrates. Journal of the North American Benthological Society, 15:185-211.
Chessman, B., Williams, S. & Besley, C. 2007. Bioassesment of streams with
macroinvertebrates: effects of sampled habitat and taxonomic resolution. Journal of the
North American Benthological Society, 26:546-565.
Cota, L., Goulart, M., Moreno, P & Callisto, M. 2002. Rapid assessment of river quality using
an adapted BMWP index: a practical tool to evaluate ecosystem health. Verh. Internat.
Verein. Limnol., 28:1-4.
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Costa, J. M., Souza, L. O. I. de & Oldrini, B. B. 2004. Chave para identificação das famílias e
gêneros das larvas conhecidas de Odonata do Brasil: comentários e registros
bibliográficos (Insecta, Odonata). Publicações Avusas do. Museu Nacional, 99:1-44.
Fernandez, H. R. & Dominguez, E. (eds). 2001. Guia para la determinacion de los
artropodos bentônicos sudamericanos. Universidad Nacional de Tucumán. Argentina.
282pp.
Junqueira, V. M. & Campos, S. C. M. 1998. Adaptation of the BMWP method for water
quality evaluation to Rio das Velhas watershed (Minas Gerais, Brazil). Acta Limnologica
Brasiliensia, 10:125-135.
Junqueira, V. M., Amarante, M. C., Dias, C. F. S. & França, E. S. 2000. Biomonitoramento da
qualidade das águas da Bacia do Alto Rio das Velhas (MG/Brasil) através de
macroinvertebrados. Acta Limnologica Brasiliensia, 12:73-87.
Marchant, R. 2007. The use of taxonomic distinctness to assess environmental disturbance of
insects communities from running waters. Freshwater Biology, 52:1634-1645.
Melo, A. S. 2005. Effects of taxonomic and numeric resolution on the ability to detect
ecological patterns at a local scale using stream macroinvertebrates. Archiv für
Hydrobiologie, 164:309-323.
Miserendino, M. L. & Pizzolón, L. A. 1999. rapid assessment of river water quality using
macroinvertebrates: a family level biotic index for the Patagonic Andean zone. Acta
Limnologica Brasiliensia, 1:137-148.
Olifers, M. H., Dorvillé, L. F. M., Nessimian, J. L. & Hamada, N. 2004. A key to Brazilian
genera of Plecoptera (Insecta) based on nymphs. Zootaxa, 651:1-15.
Pes, A. M. O., Hamada, N. & Nessimian, J. L. 2005. Chave de identificação de larvas para
famílias e gêneros de Trichoptera (Insecta) da Amazônia Central, Brasil. Revista
Brasileira de Entomologia, 49:180-204.
Rosenberg, D. M. & Resh, V. H (eds). 1993. Freshwater biomonitoring and benthic
macroinvertebrates. Chapman & Hall, New York. 438pp.
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Sickle, J. V., Larsen, D. P. & Hawkins, C. P. 2007. Exclusion of rare affects performance of
the O/E index in bioassessments. Journal of the North American Benthological Society,
26:319-31.
Stribling, J. B., Jessup, B. K. & Feldman, D. L. 2008. Precision of benthic macroinvertebrates
indicators of stream condition in Montana. Journal of the North American Benthological
Society, 27:58-67.
Whitier, T. R., Stoddard, J. L., Larsen, D. P. & Herlihy, A. T. 2007. Selecting reference sites
for stream biological assessments: best professional judgment or objective criteria.
Journal of the North American Benthological Society, 26:349-360.
Zweigg, L. D. & Rabeni, C. F. 2001. Biomonitoring for deposited sediment using benthic
invertebrates: a test on 4 Missouri streams. Journal of the North American Benthological
Society, 20:643-657.
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Os resultados obtidos nesse estudo indicam que a decomposição foliar pode ser um
indicador de qualidade ambiental em sistemas lóticos de ordens iniciais. Cuidados, porém, são
necessários quanto ao uso desse método. Neste trabalho, a detecção de alteração ambiental em
escalas de maior abrangência, como no caso da atividade agrícola ao redor dos cursos d’água,
não foi possível, muito provavelmente, devido ao pequeno número amostral e à alta
correlação entre os pontos de atividade agrícola intensa com os ambientes próximos às
nascentes na bacia hidrográfica em estudo. Esse problema referente ao delineamento amostral
pode ter camuflado informações pertinentes ao vínculo da decomposição foliar com o
gradiente de atividades agrícolas.
Visto a atividade agrícola, ou até mesmo, a atividade urbana e industrial, apresentar
diferentes formas de alteração na estrutura e função da paisagem (e.g., aplicação de
inseticidas, enriquecimento de nutrientes na terra, alteração do pH do solo, modificações na
formação vegetacional, supressão da vegetação ripária, alteração da vazão dos cursos d’água,
soterramento da calha dos cursos d’água por material rochoso de menor granulometria, entre
outros), deve-se compreender mais profundamente o quanto cada uma dessas alterações
ambientais (mas também as alterações ambientais conjuntamente) afetam o processo
ecológico de decomposição. Algumas modificações da paisagem podem agir de modo direto
sobre a decomposição foliar, como é o caso do soterramento da calha dos rios ou da alteração
da vazão dos cursos d’água por meio de atividades de irrigação (com retirada de água
diretamente dos sistemas lóticos e não de uma bacia de contenção, por exemplo). Essas
modificações, quer sejam originadas naturalmente (como no caso da maior vazão provocada
por chuvas), quer sejam provocadas pelo uso da terra (como no caso da sedimentação), foram
detectadas como agentes no processo de decomposição foliar neste estudo.
Há, ainda, efeitos indiretos da alteração estrutural e funcional da paisagem, por meio
de atividades agrícolas ou urbano-industriais (como o enriquecimento do solo por nutrientes
ou a sua contaminação por metais pesados) que podem ocasionar alterações nesse processo
ecológico. Essas alterações podem, através da lixiviação dos nutrientes ou do carreamento de
metais pesados para os cursos d’água, atingir a fauna aquática e, com isso, alterar a
decomposição foliar. Neste estudo, não foi possível detectar esses efeitos, uma vez que a
bacia hidrográfica em questão apresentava propriedades químicas semelhantes entre os pontos
62
estudados. Sabe-se, porém, que o enriquecimento de nutrientes pode alterar a composição da
assembléia de macroinvertebrados bentônicos ou provocar modificações na assembléia de
fungos decompositores. A contaminação por metais pesados pode, do mesmo modo, interferir
nesses grupos aquáticos, por meio da diminuição da abundância ou riqueza dos diversos
grupos presentes. Esses efeitos sobre a fauna podem, assim, modificar a taxa com que o
material foliar é decomposto.
A mudança na estrutura da vegetação ripária verificada neste trabalho alterou, em
primeiro lugar, os processos de deposição de sedimento, oriundos da margem do rio, e
ocasionou um retardamento no processo de decomposição foliar. Partículas de menor
granulometria, como silte e argila, apareceram relacionadas com o retardamento da
decomposição foliar, ao passo que partículas de maior granulometria - no caso, a areia -
estavam relacionadas aos trechos com maiores velocidades de decomposição. Ainda que a
areia seja uma textura de pequena granulometria, a sua presença estava negativamente
relacionada com a presença de silte e argila, que apresentam texturas ainda menores.
Trabalhos que focam esse aspecto da sedimentação têm verificado a influência das partículas
de menor granulometria como agentes do retardamento da decomposição foliar.
Outras mudanças na estrutura da vegetação ripária, como o corte e supressão dessa
vegetação, podem acarretar, ao mesmo tempo, um aumento do aporte aluvial devido à não-
retenção da água das chuvas pelo que antes seria a mata ciliar. Esse fato pode provocar (i)
uma maior perda de biomassa no material orgânico alóctone, proveniente de trechos
superiores do rio, retido em remansos ou corredeiras ou ocasionar, até mesmo, (ii) um
completo carreamento desse material em direção às partes baixas da bacia hidrográfica. Nos
locais onde o material orgânico alóctone é carreado devido ao aumento da vazão, há perda de
matéria e energia responsáveis por grande parte da produção primária. Já nos trechos onde a
biomassa vegetal apenas se degrada mais rapidamente devido ao maior volume de água
circulante, há mudanças estruturais e funcionais em suas assembléias aquáticas, o que afetará,
conseqüentemente, a cadeia alimentar envolvida.
Mudanças na composição da vegetação ripária, em termos de abundância e riqueza da
flora, podem, também, ser importantes na modificação do processo de decomposição foliar.
Neste trabalho, utilizou-se a mesma espécie vegetal em todos os pontos amostrados, o que
excluiu a possibilidade de haver diferenças químicas e físicas quando se trata do material
foliar exposto nos cursos d’água. Contudo, a exposição de folhas com alta taxa de
63
decomposição, ricas em nutrientes, misturadas com folhas de espécies pobres quanto aos
constituintes químicos, e, por isso, com baixa taxa de decomposição, podem alterar o processo
de decomposição foliar das espécies envolvidas. Isso ocorre porque os nutrientes da espécie
que se degrada mais rapidamente podem modificar a palatabilidade do material mais pobre
nutricionalmente, e fazer com que essa última espécie apresente uma decomposição mais
acelerada. O efeito dessa diversidade foliar na decomposição está, contudo, sendo recém
estudado, e as conseqüências dessa mistura de espécies são, ainda, desconhecidas.
Ao contrário do que acontece em sistemas temperados, contudo, a decomposição foliar
em sistemas tropicais ou sub-tropicais parece mais suscetível a ações físicas, como a
deposição de areia ou a alta velocidade da água, mencionadas anteriormente, e menos
suscetível a ações biológicas, ao menos no que se refere aos macroinvertebrados bentônicos.
Enquanto que a maior deposição de silte e de argila afeta negativamente a decomposição
foliar por soterramento do material foliar, o aumento da vazão influencia positivamente esse
processo ecológico. Esse aumento na velocidade de decomposição, por meio do aumento da
vazão, gera maiores efeitos sobre a fauna bentônica que mais diretamente "depende" do
recurso foliar, a saber, os organismos retalhadores. Quando a velocidade da água diminui, e
assim, a decomposição foliar desacelera, a fauna bentônica responde positivamente à presença
prolongada do recurso alimentar. Os organismos do tipo funcional retalhador podem, então,
ocorrer sobre esse material e retirar os nutrientes necessários para seu crescimento. Sem a
presença tão constante do recurso foliar, tais organismos retalhadores não poderão adquirir a
matéria e energia necessárias para seu crescimento a partir do material orgânico alóctone.
Além disso, pressões ecológicas de cima para baixo na cadeia alimentar (relação top-down) -
nomeadamente a predação, a qual é considerada como muito importante no controle da fauna
em sistemas tropicais e subtropicais - fazem com que a fauna detritívora apresente dieta mais
generalista para obtenção de alimentos do que o demonstrado para sistemas temperados, onde
essa fauna pode ser especificamente retalhadora de folhiço, sem estar sujeita à pressão
exercida pelos predadores.
O fato de sistemas neotropicais não apresentarem quantidade idêntica de retalhadores
em relação aos sistemas temperados parece estar relacionado, assim, aos motivos hidro-
metereológicos, típicos de sistemas tropicais e sub-tropicais (i.e., chuvas intensas e
freqüentes), que não permitem a manutenção tão prolongada do recurso foliar nos cursos
d’água, comum nas regiões mais frias e menos chuvosas do hemisfério norte. A pressão
ecológica por parte dos predadores, contudo, não foi verificada neste trabalho, uma vez que a
64
razão entre predadores e presas permaneceu a mesma nos dois tratamentos (com alta e baixa
vazão). Sugere-se que a pressão por predação seja, de qualquer modo, importante na
regulação da ocorrência de grupos detritívoros com hábitos específicos em regiões tropicais e
subtropicais.
Desse modo, a hipótese de que o processo de decomposição foliar seja afetado pela
poluição, tanto urbana quanto agrícola, através de mudanças estruturais e funcionais na
assembléia de macroinvertebrados aquáticos que, por sua vez, ocasionaria modificações no
modo como o material foliar é consumido, não é sustentada a partir dos resultados obtidos no
presente estudo. A poluição por atividades agrícolas parece ocasionar efeitos mais intensos na
decomposição através de alterações físicas do ambiente do que através de alterações
biológicas. Contudo, mesmo que o aumento no fluxo da água seja responsável por aumento na
decomposição, não se sabe, ainda, o quanto os organismos retalhadores realmente atuam na
decomposição (ou seja, o quanto eles decompõem o material foliar), nem o quanto outros
organismos microbiológicos, como os fungos aquáticos, são importantes para o processo de
decomposição. Visto, porém, que os retalhadores não aparecem como principais causadores
da decomposição foliar, ao contrário das regiões temperadas, talvez sejam os fungos aquáticos
os principais agentes biológicos envolvidos no consumo do material foliar.
Este trabalho também indicou que todos os pontos amostrados se enquadram em
categorias de risco ambiental com baixa qualidade das águas, quando a fauna aquática é
analisada. A partir do uso de índices BMWP com valores da Patagonia argentina e do Cerrado
brasileiro, obtiveram-se maiores valores biológicos, isto é, melhor qualidade de água,
justamente nos mesmos trechos que apresentaram maior quantidade de retalhadores. O fato de
os retalhadores se apresentarem como grupo funcional indicador de qualidade ambiental
explica-se devido ao vínculo desse grupo de invertebrados com o material alóctone e, assim,
com a presença de mata ciliar em melhor estado de conservação. Porém, estudos mais amplos
sobre a fauna bentônica e suas respostas frente a alterações ambientais são necessários antes
que esses organismos possam ser aplicados integralmente no continente latino-americano, em
especial, no sul do Brasil, onde ainda há escassez de informações referentes à ocorrência dos
macroinvertebrados bentônicos.
O processo ecológico de decomposição foliar, influenciado mais fortemente por
fatores físicos ou biológicos, é, ainda, de grande importância para os sistemas aquáticos
continentais. Esse processo transforma a matéria orgânica particulada grossa, consumível por
65
poucos organismos, em matéria orgânica particulada fina, a qual é a base da cadeia alimentar
detritívora em sistemas aquáticos. Entender como esse processo ocorre em sistemas tropicais
e subtropicais e compreender as conseqüências que as atividades humanas trazem para a
decomposição desse material orgânico, é peça fundamental para melhorar a compreensão
ecológica dos sistemas hídricos (em especial, daqueles de água corrente) e, a partir daí,
estabelecer critérios ambientais quanto ao uso das terras em bacias hidrográficas. Assim, para
que as políticas públicas de uso dos cursos d’água em países como o Brasil sejam
desenvolvidas visando à conservação da funcionalidade desses ambientes, deve-se,
primeiramente, saber como esses sistemas aquáticos realmente funcionam em termos
ecológicos e hidrológicos.
66
R
R
E
E
F
F
E
E
R
R
Ê
Ê
N
N
C
C
I
I
A
A
S
S
B
B
I
I
B
B
L
L
I
I
O
O
G
G
R
R
Á
Á
F
F
I
I
C
C
A
A
S
S
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73
A
A
N
N
E
E
X
X
O
O
S
S
Tabela 1. Tabela filogenética do total de amostras realizadas na Bacia do Lajeado Grande, de
novembro de 2004 a julho de 2005.
Classe
Ordem Família Gênero Nº de indivíduos
Insecta Trichoptera Calamoceratidae Phylloicus 19
Hydropsychidae Smicridea 128
Hydroptilidae Neotrichia 34
Hydroptila 64
Leptoceridae Nectopsyche 350
Oecetis 2
Policentropodidae Cernotina 6
Coleoptera Elmidae Elm_sp1 668
Elm_sp2 5
Elm_sp3 4
Elm_sp4 10
Elm_sp5 1
Dytiscidae Dyt_sp1 7
Noteridae Not_sp1 3
Gyrinidae Gyr_sp1 24
Ephemeroptera
Leptophlebiidae Ulmeritoides 173
Miroculis 40
Simothraulopsis 2
Caenidae Caenis 399
Leptohyphidae Traverhyphes 452
Lepto_sp1 7
Trichorythodes 4
Baetidae Bae_sp1 31
Plecoptera Perlidae Anacroneuria 10
Gripopterygidae Limnoperla 56
Zygoptera Calopterygidae Hetaerina 23
Coenagrionidae Acanthagrion 6
Coen_sp1 6
Coen_sp2 43
Megapodagrionidae
Allopodagrion 23
Protoneuridae Peristicta 7
Anisoptera Libellulidae Lib_sp1 2
Lib_sp2 4
Cordulidae Navicordulia 3
Cord_sp2 3
Gomphidae Cacoides 4
Phyllocycla 7
Aeshnidae Aesh_sp1 1
Lepidoptera Pyralidae Nymphulinae 1
Lep sp1 Lep_sp1 1
Diptera Chironomidae 5120
74
Simulidae 541
Ceratopogonidae 73
Tipulidae 38
Hemiptera 10
Crustacea Decapoda Trichodactylidae Trichodactylus. 11
Aeglidae Aegla 2
Annelida Oligochaeta 326
Hirudinea 86
Mollusca Gastropoda Ampullariidae Pomacea 3
Ancylidae Gundlachia 14
Drepatonema 5
Hydrobiidae Potamolithus 15
Heleobia 1
Planorbidae Drepanotrema 4
Bivalvia Sphaeriidae Pisidium 95
Nematyelminthes 2
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